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)

property. The monitoring wells range in depth from 20 to 25 feet bgs, and are sampled on a semiannual
basis. No soil or groundwater sampling appears to have been conducted off-site. Therefore, the extent of
contamination in both of the horizontal and vertical directions, does not seem to be delineated.
Sampling of the monitoring wells showed highest concentrations of contaminants in MW-1 (next to the
gasoline USTs). Concentrations in MW-1 dropped significantly over the sampling period on record, with
MTBE concentrations dropping from 210,000 ppb in 4/2006 to 54 ppb in 3/2011. Similarly, benzene
concentrations dropped from 1,450 ppb to 55 ppb during the same time period. In contrast, concentrations
of ethyl benzene and total xylenes did not change as significantly during the sampling period. For example,
ethyl benzene concentration measured at 3,610 ppb in 4/2006 compared to 3,840 ppb in 3/2011. Total
xylenes on the other hand, were measured at 18,600 ppb in 4/2006, compared to 12,300 ppb in 3/201112.
This would seem to indicate that MTBE and benzene were quick to move with the groundwater and away
from the site. This could be attributed to the transport properties of the different chemicals, with MTBE and
benzene being more mobile and less likely to adsorb to soil media. Ethyl benzene and xylenes on the other
hand have higher retardation coefficients compared to benzene and MTBE.
Modeling work conducted by the consultants (MTBE plumes) indicates that the MTBE plume likely
moved off site. A potable well, located on the northwest end of the property did not show contamination
over the period of sampling from 2006 to 2012.

4.3 TS#3 Getty West Windsor
This site is located at 3710 Route 1, Princeton, in Mercer County. General topography around the area is
relatively flat, and land use near the gasoline station is mostly commercial and residential. Approximately
927 feet north of the service station is an “unnamed body of water” (Tyree, 2006). The Millstone River and
Carnegie Lake are approximately 3,333 feet northeast, and 3,899 feet north of the service station
respectively.
Geology and Hydrogeology: The site is located within the Lockatong Formation which consists of “black
and gray siltstone and shale ranging from thin to thick layers” (Tyree, 2006). Based on soil borings
collected at the site, consultants describe the subsurface matrix as “red brown silts and fine to coarse sands
some clays to approximately 20-feet below ground surface (bgs). Red brown fractured shale was
encountered to approximately 46-feet bgs. Competent bedrock was encountered from 46-feet to
approximately 61-feet bgs.” (Tyree, 2006).
Groundwater at the site is between 13 to 21 feet bgs, and flows to the east-southeast. The average hydraulic
gradient is estimated at 0.0194ft/ft, the hydraulic conductivity (K) has been calculated to be 3.8 ft/day, and
the groundwater velocity was calculated at 69.35 ft/year (0.19 ft/day) (Tyree, 2006).
Dissolved oxygen concentrations in the monitoring wells, measured in January 2006, were below 1 mg/l in
all wells, and the pH ranged between 6.43 and 7.16.
12

Bates stamp # SOVCON015058
37

Nearby receptors: A well records search with the Bureau of Water Allocation (BWA) identified 7 domestic,
2 irrigation, 3 industrial, and 3 public wells within a one mile radius of the site. A door-to-door canvas
identified three potable domestic wells within a 1000 feet radius, all located hydraulically down gradient
from the service station (Tyree, 2006).
Release History: In July of 2004, a product line test failed, resulting in the initiation of soil investigation
activity at the site. Results indicated BTEX and MTBE contamination in exceedance of the state’s Soil
Cleanup Criteria (Criteria). In June and July of 2005, USTs and associated piping were removed from the
site. Excavated soils, approximately 5,600 tons, removed from the area of the unleaded gasoline UST and
the dispenser island, were taken offsite for thermal destruction. Soil analysis conducted in August of 2004
showed significant MTBE contamination in 7 out of 19 soil borings examined (concentrations ranging from
10 to 696 mg-MTBE/kg-soil, well above the State’s Criteria of 3.1 mg/kg). In contrast, Benzene
concentrations were above the State’s Criteria of 1 mg/kg in 3 out 19 soil borings examined.
Concentrations ranged between 2.77 to 4.78 mg-benzene/kg-soil.
Monitoring and Remediation: Between the end of 2004 and the beginning of 2006, twelve monitoring wells
were installed onsite. All monitoring wells had relatively high concentrations of benzene and MTBE
indicating the plume was not delineated, and was potentially moving offsite. Four additional monitoring
wells were later installed on an adjacent property across the street and in the general down gradient
direction from the contamination site. MTBE and TBA concentrations measured in those four wells were
mostly below the State’s GWQS (70 ppb). However, given that the wells are shallow, it is probable that the
MTBE/TBA plume moved below the monitoring network (see Ellis 2000; API 2000; Wilson et al 2005c;
API, 2006).
It should be noted that the maximum MTBE and TBA concentrations measured at the site are considerably
high. For example, maximum MTBE concentration was measured in MW-5 (converted into an extraction
well in early 2006) at 2,030,000 ppb (2,030 ppm) on 11/10/2004. The maximum TBA concentration for the
site was measured in the same well (MW-5) at 333,000 ppb (on 5/12/2005). In contrast, the maximum
benzene concentration was recorded in well MW-6 (converted to extraction well in early 2006) at 11,300
ppb, on 3/1/2005.
Remedial action has been taken at the site. Four groundwater extraction wells were installed, and by April
2010 an estimated 462,111 gallons of groundwater had been extracted (Tyree, 2010). Additionally, a soil
vapor extraction system (SVE) was to be installed at the site to augment site remediation (Tyree, 2010).

4.4 TS#4 Exxon Livingston
The site is located at 38 East Mount Pleasant Avenue, Livingston, in Essex County. The gasoline station is
in an area with commercial and residential property. The site slopes to the west along East Mount Pleasant
Avenue. The surface water body nearest the site is Canoe Brook, located approximately 950 feet southwest
of the site.

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Geology and Hydrogeology: Subsurface activities at the site indicate that the overburden consists of
“brown silt with fine sand up to an approximate depth of 30 feet below grade and brown fine to coarse sand
with well-rounded coarse gravel up to an approximate depth of 40 feet below grade” (Kleinfelder 2010).
This is characteristic of the Pleistocene glaciation in the area. The consultants also note that: “the sand and
gravels tend to sit within the buried valleys of the bedrock topographic surface and act as groundwater
recharge areas when present near the current land surface” (Kleinfelder 2010).
At the site, bedrock was encountered at depths somewhere between 20 and 45 feet bgs. The bedrock is
described as being part of the Towaco formation, and “consists of reddish brown to brownish purple, fine to
medium grained sandstone, siltstone, and silty mudstone. These deposits are generally located within
upward fining sequences one to three meters thick. The sandstones of this formation are commonly crosslaminated, and the siltstones are commonly planar laminated or bioturbated. Locally, beds in the Towaco
formation strike north 45 degrees and dip 7 to 12 degrees to the northwest in Essex County. Both vertical
and horizontal fractures are common within this formation (Drake et al., 1996)” (Kleinfelder 2010).
Groundwater in the overburden is encountered at depths of 30 to 40 feet bgs, and moves predominately to
the west. The hydraulic conductivity was calculated to be 3.805 ft/day based on testing done in June 2003.
The hydraulic gradient was estimated between 0.02 and 0.07 ft/ft (based on wells MW-1 to MW-5,
measured three different times between September 2009 and June 2010).
In the bedrock, the hydraulic head is “highly variable and ranges from approximately 30 to 112 feet below
grade” (Kleinfelder 2010). Based on aquifer testing done in January 2007 on two bedrock wells (MW-5D
and MW-11D) the flow of water in the bedrock aquifer was concluded to be “predominately along strike to
the southwest” (Kleinfelder 2010). The hydraulic conductivity in the fractures was calculated at 0.9378
ft/day, and in the bedrock matrix at 9.678x10-6 ft/day.
Nearby receptors: The closest commercial supply well is approximately 800 ft down gradient (southwest)
of the site. During sampling events in October 2009 and April 2010, MTBE was detected at concentrations
of 0.13 and 0.19 ppb respectively. No BTEX were detected in the well.
Release History: The site contains three fiberglass gasoline USTs (8,000, 10,000 and 12,000-gallons) which
were installed in 1986 (Kleinfelder 2010). Site investigation activity was initiated after the accidental
discharge (“drive-off”) of an estimated one-gallon of regular gasoline to the pavement and the pea gravel
beneath the dispenser island in May 2001. In February 2002, approximately 0.09 feet of liquid phase
hydrocarbon (LPH) was detected in one of three monitoring wells (MW-1) installed at the time. In March
2003, LPH was detected in two additional monitoring wells (0.17 feet in MW-4, and 0.1 feet of LPH in
MW-6). The maximum LPH was detected at 0.48ft in monitoring well MW-4 in February 2004. By March
2004, “five 1,000-gallon gasoline USTs that were previously abandoned-in-place were identified and
removed from the site” (Kleinfelder, 2010). Between June 2002 and April 2004, approximately 20 gallons
of LPH were extracted from the site as part of interim remedial activities.

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Monitoring and Remediation: The monitoring systems includes 14 overburden and 14 bedrock monitoring
wells. The consultant’s report indicates that remediation involves groundwater pump and treat (GW P&T)
and soil vapor extraction (SVE), and the monitoring network is relatively large. Onsite, there are 8
extraction wells (MW-1 to MW-8) 3 bedrock monitoring wells (MW-5D, MW-5D2R and MW-5D3) and
“active GWP&T and SVE remedial systems.” Additionally, 16 monitoring wells (MW-9, MW-9D, MW10D, MW-10R, MW-11D, MW-11R, MW-12D, MW-13, MW-14S, MW-14D, MW-15S, MW-15D, MW16D, MW-16D2, MW-17D and MW-17D2) are located offsite on nearby properties. The SVE system was
started in July of 2004, and the GWP&T system was started in November of the same year.

4.5 TS#5 Sunoco Bloomfield
The site is at 72 Bloomfield Avenue and 17th street in Bloomfield, Essex County, and has operated as a
petroleum retail station since 1947.
Geology and Hydrogeology: The general area in the vicinity of the site is characterized by overburden on
top of bedrock. The bedrock, of the Passaic formation, is “overlain by approximately 50 to 100 feet of
unconsolidated glacial/fluvial deposits, consisting of silt, clay, and sand” (GES, 2006). At the site, splitspoon sampling and drill cuttings indicate the overburden consists of brown sand to depth of five to ten feet
overlying at least 15 feet of compact silt. Soil borings for the installation of monitoring wells extended only
to a depth of 25 feet bgs, and depth to bedrock was not evaluated.
The average hydraulic conductivity for the site was calculated at 0.168 ft/day (5.93x10-5 cm/sec). The
groundwater flow, based on the sampling event in November 2005, was to the northeast at an average
hydraulic gradient of 0.015 ft/ft (GES, 2006).
Nearby Receptors: A search of the records of the NJDEP Bureau of Water Allocation indicates the
presence of twelve wells (one domestic, seven commercial, and four industrial) within a one-mile radius of
the site. The domestic well is upgradient (approximately 700 ft southwest) of the site. All identified wells
are in bedrock, with depth of 110 feet or more.
Release History: In January 1992, soil contamination was discovered during the removal of a 1000-gallon
UST from the site. Consequently, approximately 28 tons of soil was excavated and removed form the site
(GES, 2006). Groundwater sampling was started in 1992.
The monitoring network includes 9 on-site, and 9 off-site monitoring wells, sampled semi-annually.
Remediation activity: groundwater extraction and SVE, were started in May and June of 2005 respectively
(GES 2006). By early 2006 approximately 26,114 gallons of groundwater were recovered. In their 2006
report, the consultants note that the remediation system does not address contamination off-site noting: “It
is acknowledged that the VOC concentrations in groundwater at offsite MW-9 are not addressed by the
existing remedial system” (GES, 2006).

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4.6 TS# 6 Hess Maple
The site is located at 2815 Route 73 and Regent Avenue, Maple Shade in Burlington County. It is located
in the Inner Coastal Plain. Some 450 feet west of the Site is a “deciduous wooded wetland” which
surrounds the South Branch Pennsauken Creek. The Creek, the surface water body nearest to the Site, is
approximately 600 feet to the west.
Geology and Hydrogeology: The site slopes gently to the southwest, and based on site investigation
activity, groundwater flows to the west-southwest, and the average hydraulic gradient is 0.012 ft/ft.
The consultant’s report (Envirotrack, 2012) describes the local geology as consisting of Upper Cretaceous
Age sedimentary rocks, adding that: “Unconsolidated Pleistocene and recent deposits form a thin,
discontinuous cover over the bedrock surface. Bedrock geology underlying the Site is the Woodbury
Formation which consists of unconsolidated clay, sand, and silt” (Envirotrack, 2012; emphasis added).
Release History: According to the consultant’s report (Envirotrack, 2012), contamination at the site was
first noted in December 1989, “during closure activities of three 8,000-gallon steel underground storage
tanks.” At the time, both “impacted soil and LNAPL were observed”. In May 1990, three 10,000-gallon
fiberglass USTs were installed. Not long after, in August 1996, “Hess ceased retail operations at the site”
and divested the property to ARFA. However, in October 1999 and April 2000, elevated concentrations of
MTBE were reported in the monitoring wells in near the UST field. As a result, the NJDEP ordered ARFA
to conduct a 7-day investigation into the source of the increased concentrations. Based on LNAPL
characterization, the investigation concluded that a discharge likely occurred between September 1999 and
August 2000 (Envirotrack, 2012).
The consultant’s report however, indicates the presence of a significant amount of LNAPL at the site over
the span of many years:
“The maximum horizontal extent of the LNAPL body was 92 feet long in the direction of
groundwater flow and 52 feet wide perpendicular to the direction of groundwater flow on
December 19, 2000. As of August 18, 2011, the horizontal extent of the LNAPL body was 50 feet
long in the direction of groundwater flow and 45 feet wide perpendicular to the direction of
groundwater flow. The vertical extent of the LNAPL body has ranged from 14.25 feet below the
top of casing (MW2 on November 27, 1992) to 8.07 feet below the top of casing (MW5 on April
30, 1993). The thickness of the LNAPL body has ranged from 0.01 feet (MW2 on December 19,
2000) to 2.80 feet (TW-N on December 19, 2000). During the August 18, 2011 groundwater
gauging and sampling event, LNAPL was detected in TW-N at a thickness of 0.21 feet, TW-E at a
thickness of 0.25 feet, TW-W at a thickness of 0.25 feet, and TW-S at a thickness of 0.12 feet.”
(Envirotrack, 2012)
The 2012 consultant report indicates the presence of LNAPL in 4 out of 27 monitoring wells. On February
1, 2012, samples of groundwater and LNAPL were collected from tank field well TW-W and sent to
Torkelson Geochemistry for analysis. Torkelson “reported that the TW-W extract LNAPL/groundwater
sample appears to be a mixture of extremely weathered gasoline and a smaller amount of severely
weathered middle distillate, either diesel fuel or fuel oil” (Envirotrack, 2012). However, “the age and the
41

octane rating of the LNAPL sample submitted to the lab could not be estimated” (Envirotrack, 2012).
As of the most recent consultant’s report: “[a]quifer testing has not been completed” and “LNAPL
mobility, recoverability or treatability tests have not been completed” (Envirotrack, 2012).
Monitoring and remediation: The monitoring wells appear to be shallow for the most part, with depth to 25
feet. Offsite wells include monitoring wells MW-18 through MW-22 in the downgradient direction.
Remediation efforts include groundwater extraction, which was conducted between November 1990 and
December 1995 from four monitoring wells (MW1, MW2, MW3, and MW5). During that time period, an
estimated “537,000 gallons of combined LNAPL and groundwater” were recovered (Envirotrack, 2012). In
1998, groundwater extraction was restarted (wells MW2 and MW5) and continued through November
2002. During this second phase, an estimated 36,000 gallons of “combined LNAPL and groundwater” were
recovered. More recently, extraction was started again in March 2011 (wells TW-N, TW-E, TW-W, and
TW-S), and continues as of the date of the most recent consultant report (Envirotrack, 2012). Additionally,
“Total-Phase Extraction13 (TPE) remediation events began in May 2011” targeting monitoring well MW11,
as an additional remediation measure.

4.7 TS#7 Gulf Waldwick
The site, an active Gulf Service Station, is located at 49 Franklin Turnpike Waldwick, Bergen County. It is
surrounded by a mixture of residential and commercial properties. The site is situated within the Piedmont
Physiographic Province.
Several surface water bodies are located within a half-mile radius of the site, and those include:

-

Cole Pond (1,350 feet south, sidegradient to site)

-

Hohokus Brook (1,600 feet south, sidegradient to site)

-

Smokis Vall Brook (1,950 feet west, sidegradient to site)

Geology and Hydrogeology: The local geology is described as consisting of “reddish-brown pebble
conglomerate, pebbly sandstone, and sandstone in upward-fining sequences three feet to six feet thick.
Clasts are sub-angular to sub-rounded quartz and quartzite in the sandstone matrix” (Brilliant Lewis, 2009).
Kleinfelder (2007) described the site geology at different depths below ground surface (bgs) as:

-

From 0 to 20 feet bgs: “brown silty sand, some gravel and clay.”

-

From 20 to 50 feet bgs: “reddish brown medium to coarse sand, minor gravel.”

-

From 50 to 70 feet bgs: “brownish red medium sand/silt, minor gravel.”

-

From 80 to 120 feet bgs: “red medium grained sandstone.”

The aquifer is unconfined, and groundwater tends to occur at a depth of 9 to 12 feet bgs (Brilliant Lewis,
13

Although not clear from the consultant’s report, TPE appears to be dual phase extraction.
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2009).
Hydrogeology: Groundwater flow velocity was calculated to be 1.7 ft/yr, based on the shallow zone (0-25
ft bgs). The hydraulic conductivity (K) was calculated to be 0.018 ft/day (for 0 – 25 ft bgs), and 0.35ft /day
for groundwater at depth 30 to 40 ft bgs. Groundwater flow direction was estimated to be to the southwest
(Kleinfelder, 2007).
For well MW-4D (Nov. 2005): Storativity (S) = 1.242 square ft/day; Transmisivity (T) = 0.2327;
Conductivity (K) = 0.018 ft/day
For well MW-16 (February 2006): Storativity (S) = 16.91 square ft/day; Transmisivity (T) = 0.01499;
Conductivity (K) = 0.242 ft/day
For well CRD-125 (July 2006) K= 0.35ft/day
Nearby Receptors: Within a one-mile radius, seven public CWS wells were identified, though none of the
wells are within the half-mile radius. The closest public CWS wells are (Kleinfelder, 2007):

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Waldwick Water Department’s Well 2, located approximately 2,700 feet
west/northwest, sidegradient.

-

Ridgewood Water Department’s Andover Well, located approximately 2,900 feet
south/southwest, downgradient.

-

Hohokus Water Department’s Wells 5 and 6, located approximately 3,280 and
4,300 feet respectively, and south/southeast, sidegradient of the site.

Release History: Contamination at the site appears to be the result of multiple release events as explained
below:

-

In July 1996, a “release of an unknown quantity of product was identified during
tank and line testing” (Brilliant Lewis, 2009). As a result, limited excavation was
conducted with approximately 200 cubic yards of impacted soil being removed
from the site. During that time, “an abandoned dispenser, associated product
piping, and a previously unknown 550-gallon heating oil UST were also
decommissioned.” (Brilliant Lewis, 2009). Nine monitoring wells were installed.

-

In February 1998, while conducting tank gauging activities, it was discovered that
a fiberglass UST (on the eastern edge of the UST field) was cracked. The amount
of gasoline released could not be identified. In response, “10 enhanced fluid
recover (EFR) events” and “15 multiphase recovery (MPR) events” were
conducted. One additional monitoring well was installed.
43

-

In September 1998, another release was reported to NJDEP after “initial tank
readings indicated that 985 gallons of gasoline had potentially been released and
subsequent stick readings were inconsistent” (Brilliant Lewis, 2009). As a result,
multiple MPR events were completed between 1999 and 2002.

In September 2002, three single walled fiberglass USTs were removed and replaced with three new double
walled fiberglass USTs. At the time “802 tons of soil/pea gravel was removed” from the site and
transported for disposal. Between 2003 and 2008 several MPR events were conducted, and eleven
monitoring wells were installed. Also, during 2008 (between January and October) ten multiphase
extraction/air sparging (MPR/AS) events were conducted (8 hours each). These events were carried out
using well MW-16 for air sparging and wells MW-4D and MW-5 for extraction (Brilliant Lewis, 2009). In
total, it was estimated that 13.6 pounds of vapor phase hydrocarbons were recovered, and 21,112 gallons of
groundwater was removed in 2008.
In communication from the State in November 2004, NJDEP noted that the extent of groundwater
contamination at the site had not been delineated, noting:
“Delineation of the groundwater contamination is not complete, as results from the Burger King
property monitoring well samples still show contamination above the Department’s GWQS.
Therefore additional delineation is required down gradient of the site” (Kleinfelder 2005)
Monitoring and remediation: The site includes a relatively large monitoring network, mostly located within
the shallow zone:
On-site:

-

In the shallow zone: Seven monitoring wells (MW-1 through MW-5, MW-8, and
RW-1).

-

In the intermediate zone: two wells (MW-4D and MW-16).

-

In the deep zone: one well (BW-1).

Off-site, and adjacent to the site:

-

In the shallow zone: Five wells (MW-10 through MW-14).

-

In the intermediate zone: One well (MW-17).

-

In the deep zone: One well (BW-2).

Off-site, on properties not adjacent to the site:

44

-

Eleven wells14 are located on nearby business properties, including Waldwick
Shopping Center, a bank, and Kaplan Cleaners.

Three additional wells (MW-6, MW-7 and MW-9), depicted on a site plan previously submitted to NJDEP
(in 2006), are reported to be sealed and located onsite, but have not been located on the property by the
more recent consultants (Brilliant Lewis). Off-site, Monitoring well MW-15 also has not been located on
the grounds.
Based on results from more recent sampling record (Brilliant Lewis, 2009), the consultant concluded that
delineation of the contamination was still not completed, and additional monitoring wells would need to be
installed: “The installation of additional wells is necessary to complete characterization and delineation of
groundwater. [Cumberland Farms Inc.] (CFI) proposes to install five off-site monitoring wells.” Two of
the proposed wells would be deep, with screen lengths in the range of 75-95 and 105-125 feet bgs.

4.8 TS#8 Shell Ridgewood
The site is located at 657 Franklin Turnpike, Ridgewood, in Bergen County. It is currently an active Shell
Service Station. The property covers approximately 32,700 square feet and sits at about 107 feet above
mean sea level.
Geology and Hydrogeology: Geology in the area consists of unconsolidated material on top of bedrock.
Stratified glacial sediments cover a depth of 25 to 30 feet, and consist of “coarse to fine sand with variable
percentages of clay, silt, gravel, and cobbles” (Handex, 1988). Underneath the glacial sediments is 8 to 10
feet of “regolith; extremely weathered pieces of bedrock ranging in size from gravel to boulders inn a
clayey-silt matrix” (Handex, 1988).
Groundwater is encountered at 20 to 25 feet, and generally moves to the south/southeast. The hydraulic
gradient was estimated at 0.006 ft/ft.
Competent bedrock, fine-grained sandstone of the Brunswick formation underlies the regolith.
Release History: The exact history of the release(s) at the site is not clear. However, according to
consultant’s report site investigation was initiated in 1987 after “two municipal wells in the area of the
service station had been taken off line due to residents complaining of odors in their tap water on June 23,
1987” (Handex, 1988). In July 1987, the MTBE concentration in one of the wells, Twinney, measured at
145 µg/L, whereas in the other well, Walthery, MTBE measured at 84 µg/L in the same month. In August
1996, a third municipal well, Paramus, was shut down because of contamination (Sovereign, 2010).
Concentrations of MTBE in the three wells fluctuated significantly as can be seen from Table 4.3 below
which lists the lowest and the highest MTBE concentrations measured in the three wells, over the time
period in which they were sampled.
14

Those wells include CFS-37, CBS-50, CRD-125, KCMW-1through KCMW-7, and KCMW8D.
45

Table 4.3: Range of measured MTBE concentrations in the three municipal wells within a halfmile radius of the Shell Ridgeway Site.

Well
Name

Period
sampled

Twinney

11/21/836/28/00

Lowest MTBE measured

Highest MTBE measured

Concentration
(µg/L)
µ

Date

Concentration
(µg/L)
µ

Date

0.49

6/28/00

197

2/24/98

Walthery 7/5/83-3/10/98

2.2

7/17/90

485

8/5/96

Paramus 8/8/83-9/7/99

0.6

5/28/99

689

9/9/96

Monitoring and Remediation: As a result of discovering MTBE contamination in the municipal wells, 20
shallow and 10 deep monitoring wells were installed in 1987 at the site. Shallow wells extended to a
maximum depth of approximately 39 feet below surface, whereas the bedrock wells extended to a
maximum of 96 feet (Handex, 1988). Separate phase product was discovered in wells W-1 and W-4 with
thickness up to 6.25 inches in August 1987 (Handex, 1988). Initial site investigation concluded that by
November 1987, the MTBE plume had migrated over 500 feet downgradient from the site. Furthermore,
“elevated levels of MTBE in the Walthery well and monitoring well D-1 (in bedrock) indicated that
contaminants had migrated from the water table aquifer to the bedrock aquifer” (Handex, 1988).
In September 1990, a pump and treat system was started, and in June 1992, free product was detected in
well 4 at a thickness of 1.9 feet.
In February 1995 a line-leak was reported in a 1,000-gallon, 89-octane regular gasoline tank. As a result,
two additional monitoring wells were installed, and in October 1995 “recovery wells were redeveloped via
surge/bail”. In August 1996, a soil vapor extraction (SVE) system was started.
In June 1998, there was a reported “Leaking flex connector discovered underneath the middle pump
island”, and in September 2002 ORC (oxygen release compound) was injected at the site, followed in
February 2004 with upgrades to the pump and treat system and the SVE system, with three additional SVE
wells installed. The Pump & Treat system was shut down in January 2009, and the SVE system was
eventually shut down in May 2009.
Over the years the monitoring network extended to some 60 wells onsite and offsite (see Sovereign 2010b),
with wells added on over time. The extent of the MTBE plume exceeded that of benzene as can be seen
from the monitoring record in wells 27, 33, 37, 53, 55, and others.

4.9 TS#9 Points BP
The site is located at 109 Delsea Drive, Sewell in Gloucester County. It is an active gasoline service station,
and has operated as such for “over 50 years” (ELM, 2012). The site lies within the Coastal Plain
Physiographic province.

46

The property sits on a topographical mound, and the area generally slopes to the northwest (ELM, 2012).
The surrounding property is mostly commercial. Immediately to the south of the site is an “active home
heating oil distribution facility” which is associated with a residential unit (Brilliant, 2011). The residential
unit is located west of the site. Also nearby are an auto dealership (to the north), a dry cleaner (to the
northeast) and an abandoned Exxon service station located to the east of the site.
Nearby Receptors: To the north and northeast of the site, beyond the commercial property, is an “extensive
wooded area with wetlands”. Immediately downgradient of the site is a “detention basin” on the property
with the auto dealership (Adorin Auto) (Brilliant, 2011). This property appears to be identified as “Dylans
RV Center (Dylans) in a different consultant report (ELM, 2012), and the detention basin is described as a
“storm water detention basin.” To the west of the site is another commercial property, Tortorice
Contractors, which also includes a storm water detention basin.
To the northwest, and beyond the Tortorice property is a “forested wetland, drainage features, and surface
water bodies” (ELM, 2012). The ditches and drainage features (which contribute to Monongahela Brook)
and which form drainage channels in the wooded area are believed to remnants of agricultural cultivation
dating to a period between the 1930’s through the 1960’s (based on aerial photography of the area from that
time period; ELM, 2012).
Geology and Hydrogeology: the subsurface soils at the site are described as “tan to grey, sandy silts to silty
sands with clay and intermittent gravel lenses” (Brilliant, 2011).
Groundwater flow is to the north, northwest direction (ELM, 2012). Groundwater occurs at a depth of
approximately 25 feet below ground surface. However, closer to the nearby surface water body (a man
made drainage ditch, and a forested wetland approximately 500 feet downgradient of the site), groundwater
occurs at depths of 1 to 6 feet below the surface (in monitoring wells MW-17 and MW-27); “Groundwater
intersects the channel bottom of the ditch and provides baseflow in the waterway under non-storm
conditions” (ELM, 2012).
Release History: the record seems to indicate that the first report of contamination at the site occurred in
October 1995, during soil investigation activities related to the sale of the site. Subsequently, in 1996,
during station upgrades and the removal of three gasoline USTs, “petroleum odors were noted in the soil”
and another case number was issued (Brilliant, 2011). In April 1998, liquid phase hydrocarbon (LPH) was
found at the water table at the site. Since then, LPH has repeatedly been detected (in different monitoring
wells, and at different times, with thickness up to 6.89ft) during sampling events. Based on consultant’s
report (ELM, 2012): “Remedial activities initiated in April 2000 have removed thousands of gallons of
LPH to date.”
Monitoring and Remediation: Based on ELM (2012), investigation activities have been ongoing at the site
since 1995. However, as noted earlier, LPH has been measured repeatedly in numerous monitoring wells
over time.

47

Between 2007 and 2009 free product was removed from the site using “hand bailing”. Since 2009 however,
Multi-Phase Recovery (MPR) events were used instead to extract vapors and liquid from the “pea gravel
within the tank field and pipe trenches” (Brilliant, 2011). The MPR events have involved wells MW-2,
MW-3, MW-4, MW-5, and MW-12. Between September 2009 and April 2011, 10 MPR events were
carried out, with a total of 4,129 pounds of vapor, and 5,386 gallons of water extracted form the wells.
In August 2009, an additional monitoring well (MW-24) was installed offsite, and in the downgradient
direction, on residential property. The well depth extended to 20 feet below surface. Remediation also
involved the use of oxygen release compound enhance natural degradation processes (ELM, 2012).
As of May 2011, two onsite wells (MW-5 and MW-12) contained measurable amounts of LPH. A letter
from NJDEP to ARFA Enterprises (dated February 2011) notes several deficiencies in the containment and
stabilization of contamination (presence of liquid phase hydrocarbon in several monitoring wells at the
site), as well a lack of delineation of the extent of contamination at the site, including in the horizontal and
vertical directions, and the extent of the free and residual product in the subsurface (Brilliant, 2011).
The ecological evaluation report by ELM (2012) notes that: “Source control is on-going and has not been
fully achieved.” The consultants also noted that contamination detected in surface water samples and
sediment samples from the nearby ditch/drainage features is the result of discharge of contaminated
groundwater from the site. ELM (2012) concluded: “The principle exposure pathway for contaminants
associated with the gas station property to contact sensitive receptors is through groundwater discharge to
the unnamed tributary” and “Discharge of contaminated groundwater to the unnamed tributary (ESA-3)
represents a complete exposure pathway between Site contaminants and environmentally sensitive
receptors” (ELM, 2012).

4.10 TS# 10 HP Delta
The H.P. Delta, also known as Rob’s Service Station is located at 439 Lake Avenue Woodbridge, in
Middlesex County. The property has an active gasoline station, as well as an auto repair service station
(Louis Berger, 2010). The majority of the property in the surrounding area is residential, but the property
near the intersection of Lake Avenue and Lancaster Road is mainly commercial.
The site is approximately 80 feet above mean sea level (msl) and gently slopes to the southwest. Drainage
from the site follows a west-southwest direction, towards Pumpkin Patch Brook, which is approximately
0.3 miles from the site (Louis Berger, 2010).
It is noted that the area this site is in receives an average of 44 inches in precipitation, with the heaviest
amounts normally falling in the summer months (Louis Berger, 2010).
Geology and Hydrogeology: The geology around the site is described as being “classified as Haledon
Urban Land Complex with 0-8 percent slopes. The Haledon is described as coarse loamy basal till
originally derived from basalt. Urban land is characterized by surfaces covered by pavement, concrete,
buildings, and other structures underlain by disturbed and natural soil material” (Louis Berger, 2010). The
48

till is reported to be Rahway Till, some 20 feet thick near the site, and described as consisting of “reddishbrown clayey, silty sand to clayey, sandy silt” (Louis Berger, 2010).
The underlying bedrock is the Passaic Formation. The Passaic Formation “consists of reddish-brown to
brownish-purple to grayish-red siltstone and shale that strikes northeast and dips approximately 11o
northwest in the area of the site.” The fractured bedrock is reported to occur at approximately 17 to 22 feet
below ground surface (bgs) (Louis Berger, 2010).
Groundwater in the overburden occurs at depths of 10 to 18 feet below the surface, and tends to flow in a
southwesterly direction, towards Pumpkin Patch Brook. Based on sampling done in October and December
2009, as well as in March 2010 groundwater flow in the overburden was reported to be to the west.
Fractured bedrock appears to occur at a depth of 17 to 22 feet bgs. Based on sampling done in three
bedrock wells in October 2009, groundwater flow in the bedrock was reported to be to the south, whereas
based on the sampling event done in December 2009, flow was reported to be to the southeast. A third
sampling event done in March 2010 indicated groundwater flow in the bedrock was to the south (Louis
Berger, 2010).
The local aquifer is classified as Class II-A, with groundwater that can be consumed untreated. The supply
wells “tapping this aquifer have highly variable yields reported to average 19 gmp” (Louis Berger, 2010).
Groundwater flow in the aquifer is affected by local pumping conditions as explained in the consultant’s
report:
“Under natural conditions, the regional groundwater within the underlying bedrock of the Passaic
Formation would be expected to flow northeast, toward the Rahway River, as this is a regional
groundwater discharge feature… However, the pumping of groundwater supply wells in the area
could locally affect the groundwater flow. The bedrock of the Passaic Formation is generally
considered non-porous; however, the bedrock is typically fractured along bedding planes and
contains variably-spaced joints, which tend to produce a relatively high secondary permeability.
These structural elements tend to create anisotropic flow conditions when influenced by well
pumping” (Louis Berger, 2010).
Nearby receptors: There are many domestic and irrigation wells within a one-mile radius of the site, but no
public supply wells. A memo sent from the Office of Wellfield Remediation to the Office of Community
Relations (NJDEP) dated April 28, 2009 (Louis Berger, 2010), notes the contamination of multiple nearby
residential wells with MTBE and benzene. The memo notes that sampling done by the NJDEP and the
Woodbridge Township Health Department in 2007-2009 indicated that a total of sixteen private well and
one irrigation well had detections of MTBE, benzene, and trichloroethane at concentrations above the Safe
Drinking Water Standards. The “levels of MTBE rang[ed] from 70 ppb to 12,000 ppb” (Louis Berger,
2010). The wells were reported to be mostly shallow, less than 30 feet in depth. The contaminated wells
were sealed and “a public water line was installed to supply the affected residences.” However, “it is
suspected that many residences in the area are still utilizing private wells” (Louis Berger, 2010).
Release History: The property associated with the site is owned by Mr. R. Mellecci, who was the sole
49

owner and operator of the gasoline service station until 1988. At that time (1988), Mr. Mellecci leased the
retail gasoline operation to H.P. Delta, and continued to operate the auto repair business (Louis Berger,
2010). Though no record appears to have been filed, it is believed that as part of the lease transaction, the
older USTs associated with the gasoline retail operation were replaced with the tanks currently at the site,
sometime in 1988 (Louis Berger, 2010).
The first record of a release occurred in December 2004 when “a turbine spill containment sump on the
regular gasoline UST was found to be filled with product” (Louis Berger, 2010). In the summer of 2006 the
“Woodbridge Township Health Department conducts a groundwater investigation after contamination was
detected in nearby residential potable well to the south and west of the Site” (Louis Berger, 2010). Testing
showed that 13 residences had GWQS exceedances for MTBE, TBA, and BTEX.
In February 2007, the NJDEP rejects a Site Investigation Report submitted on behalf of H.P. Delta (by
MIG Environmental) because it “did not meet minimum standards for review.” Consequently, NJDEP
conducted “publicly-funded remediation activities at the site” in the summer of 2007. Remediation
included pump and treat, as well as soil vapor extraction.
In total, there are 10 shallow monitoring wells, and three bedrock wells (to a depth of approximately 60 to
64 feet). Groundwater sampling of the monitoring wells done in October and December 2009 indicated no
contamination (benzene, TBA or MTBE) above the State’s GWQS (1, 70, and 100

g/L respectively).

However, a sampling event conducted in March 2010, two weeks after the remediation system was shut
down (pump & treat and SVE) shows exceedances for all three contaminants in MW01. Benzene, MTBE
and TBA concentrations in MW01 (a bedrock well, offsite and to the west of the Site) in March 2010 were
4.4, 120 and 740

g/L respectively (Louis Berger, 2010).

4.11 Summary
Based on information gathered at the trial sites, it is clear that MTBE has impacted the underlying aquifers
at each of the ten sites, and has impacted drinking water supplies at several of them. Remediation
(primarily groundwater extraction and or soil vapor extraction employed at some of the sites) can remove
some of the contaminant mass near the release source. However, contamination mass not captured by
remediation is free to travel with groundwater and is likely to impact other nearby receptors as it travels
away from the source.
Without significant biodegradation, MTBE attenuation in groundwater is generally poor (see Chapter 6 for
details). To my knowledge, there was no documented evidence of biodegradation of MTBE or TBA at any
of the ten Trial Sites in New Jersey. Without biodegradation or remediation, MTBE (and TBA) mass will
persist and continue to move with the groundwater to impact large volumes of water for many years to
come.

50

5

Fate and Transport of MTBE in Groundwater

In the United States today, there are thousands of public drinking water wells (and many more private
wells) in the vicinity of MTBE plumes from leaking underground storage tanks15 (LUSTs) (see Figures 2.2,
2.3 and 2.5). These “… LUST sources can persist for decades and … it can take tens to hundreds of years
for groundwater to flow from source areas to a [community water supply] CWS well” (Johnson et al.,
2000). Knowledge of the hydrogeologic processes governing MTBE fate and transport is therefore key to
assessing past, present, and potential MTBE impacts on groundwater.

5.1

Chemical and Physical Characteristics of MTBE in Groundwater

MTBE is a problematic groundwater contaminant because it is both a health concern and has
an objectionable taste and odor at very low concentrations; it is highly mobile, migrating at
nearly the same speed as the groundwater itself; extremely soluble leading to high source
concentrations at release sites; and very persistent, giving it time to spread and contaminate
large volumes of groundwater.
As of 2003, 42 states had, or were expected to have, MTBE action levels, cleanup levels, or
drinking water standards (NEIWPCC, 2003). Minute amounts of MTBE-containing gasoline
can pollute large volumes of water to the point where that water is no longer usable. “Very
small releases of MTBE (even small overfills seeping into cracks in the pavement) have the
potential to adversely impact groundwater” (Stanley, 1998). To put it in perspective, just 1
tablespoon of MTBE can render more than 586,000 gallons (2,220,000 liters) of water undrinkable (at
5 µg/l). Thus, the MTBE in reformulated gasoline16 in a 16-gallon (60.6 liters) tank of a car has the
potential to render more than 230 million gallons (870 million liters) of water undrinkable.
In effect, small volumes of gasoline can result in contamination of large volumes of water with MTBE.
Examples of case studies, documented by the State of Maine, where small surface gasoline spills resulted in
substantial contamination of groundwater wells are provided in the report by the Northeast States for
Coordinated Air Use Management (NESCAUM, 1999). In one example the report states:
“In December 1997, a car accident in the town of Standish, ME, spilled 8 to 10 gallons of gasoline
that contaminated 24 area wells with MTBE. Eleven wells were contaminated above the State’s
35µg/L (ppb [parts per billion]) state drinking water standard for MTBE. Two of the wells were
15

Most LUST sites are leaking underground fuel (gasoline) tanks (LUFTs). For the purposes of this
document, LUST and LUFT are generally interchangeable. The term LUST is commonly used in California
because of the California State Water Resources Control Board Underground Storage Tank Program.
16
RFG contained approximately 11% MTBE by volume in gasoline sold in California prior to the bans
effective on January 1, 2004
51

contaminated at levels of MTBE exceeding 1,000 µg/L, with the highest level reaching 6,500 µg/L in
the well nearest the accident site. When the State discovered the contaminated wells in May 1998, it
located and removed 79 cubic yards of contaminated soil. The contamination extended to the top of
the underlying bedrock at a depth of nine feet below the surface. The degree of soil contamination at
the bedrock surface was 100 to 230 µg/L of gasoline (total hydrocarbons).”
MTBE is normally a chemically stable compound that does not biodegrade readily in either aerobic or
anaerobic environments (Jensen and Arvin, 1990; Suflita and Mormile, 1993; Mormile et al., 1994; Yeh
and Novak, 1994; Kay-Shoemake and Watwood, 1996; Pasteris et al., 2002; Hohener et al., 2003;
Prenafeta-Boldu et al., 2004; Kim and Engesser, 2004;McClay et al., 2007; Henderson et al., 2008; Waul
et al., 2008). A detailed discussion of the microbial degradation of MTBE is presented in the following
sections.
The fate and transport properties of MTBE in groundwater are often compared to those of BTEX
components of gasoline (Benzene, Toluene, Ethyl-benzene, and Xylenes). Although conventional (nonoxygenated) gasoline can contain as many as 180 chemicals (Cline et al., 1991), the BTEX compounds
make up a relatively large percentage of gasoline (Hartley and Englande, 1992). Further, prior to the
widespread use of MTBE, the BTEX compounds were generally considered the most water-soluble
compounds contained in gasoline. Hence, the BTEX compounds typically have been the groundwater
contaminants of concern: “The components of highest aqueous concentration are the constituents of
primary concern for regulation of gasoline releases to groundwater, i.e., benzene, toluene, ethyl-benzene
and xylenes (BTEX)” (Cline et al., 1991).
However, compared to many other common groundwater contaminants including, among others, BTEX,
MTBE is more highly soluble and will therefore tend to create high concentration plumes in groundwater.
MTBE solubility17 (expressed as a concentration in water) from gasoline (at approximately 11% by
volume) is about 350 times higher than benzene, 100 times higher than BTEX, 40 times higher than
perchloroethylene (PCE) (Imhoff et al., 1995), and 6 times higher than trichloroethylene (TCE) (Russell et
al., 1992) (see Figure 5.1). The resulting MTBE concentrations in groundwater can be 10’s to 100’s of
times greater than those of benzene. Furthermore, MTBE is highly mobile in groundwater. The rate of
transport of a contaminant in groundwater can be attenuated by sorption to soil and sediments. The
octanol/water partitioning coefficient (Kow) of MTBE, which is indicative of its ability to adsorb to soil
organic matter, is one order of magnitude less than that of BTEX. As such, MTBE has been observed to
move in groundwater at about the same rate as the conservative (nonsorbing) tracer chloride (Barker et al.,
1990).
Of the BTEX compounds, benzene is the most toxic, most soluble, and most mobile. The fate of MTBE is
often compared to that of benzene. However, benzene constitutes less than 1% of the volume of gasoline,
whereas MTBE can account for up to 15% by volume of gasoline.
17

MTBE and Benzene solubilities computed for MTBE 87 octane gasoline from EPA website

http://www.epa.gov/athens/learn2model/part-two/onsite/es.htm. Last accessed 6/24/10.
52

At the same time, the fuel-water partitioning coefficient for MTBE (350) is 22 times higher than that of
benzene (15.5) (Cline et al., 1991) meaning that it is much more likely to partition out of gasoline into
water as compared to benzene. The ability to adsorb to soil organic matter is slightly higher for benzene,
compared to MTBE (ITRC, 2005), allowing MTBE to move at rates closer to those of natural groundwater
flow. Most importantly, research has shown that benzene degrades readily under aerobic conditions
(USEPA, 2005), whereas MTBE requires long acclimation times, if it degrades at all (see following
section). Under anaerobic conditions, benzene biodegradation is slower and less reliable (NRC, 2000).
Were benzene biodegradation to be limited to anaerobic conditions, the outlook for benzene natural
attenuation would not be as optimistic.

Figure 5.1. Solubility expressed as an equilibrium concentration in groundwater for MTBE from gasoline
(10.4% mass fraction), benzene (0.7% mass fraction) from gasoline, Total BTEX (9.8% mass fraction),
PCE and TCE demonstrating the potential for reformulated gasoline with MTBE to create relatively highconcentration MTBE plumes as compared to many other contaminants.

5.2 Resistance of MTBE to Biodegradation under Natural
Groundwater Conditions
Several characteristics of MTBE biodegradation, documented through observations from experimentation
and research carried out for many years now, point to the poor biodegradability of MTBE under natural
conditions. For example, early studies on MTBE biodegradability were mostly negative, and by 1998 most
published reports on MTBE biodegradation referred to the compound as recalcitrant or resistant to
biodegradation. “If a research investigation determines that a compound does not degrade, a half-life is not
reported and the compound is simply classified as recalcitrant. MTBE is generally reported as recalcitrant
and there are no widely accepted estimates of the half-life.” (USGS Fact Sheet FS-203-96; Squillace et al.,
1998).
The resistance of MTBE to biodegradation has been linked to two aspects in particular in its molecular
structure (see Figure 5.2): the tertiary carbon structure and the ether linkage (C-O-C) (Suflita and Mormile

53

1993; Salanitro 1995; Fayolle et al., 2001, Muller et al., 2007). The ether linkage is common to many
2007).
natural (e.g. lignin) and xenobiotic (e.g., agrochemicals and detergents) compounds (White et al., 1996). It
has been recognized as a “major barrier to biodegradation” since the mid 1960’s. Yet, after three decades of
research and new discoveries of organisms and pathways that can break down chemicals containing the
chemicals
ether bond, a review of many of these studies concluded:

“The ether linkage is the single most

common and unifying structural feature
which confers to both biological and
xenobiotic compounds a high degree of

resistance to biological mineralization.”
(White et al., 1996).

Figure 5.2: Molecular structure of MTBE
(from Zogorski et al., 2006).
In recent years, some studies have shown successful biodegradation under certain laboratory conditions,

while others have claimed biodegradation under natural field conditions. Meanwhile, however, the results
of numerous other scientific studies confirmed MTBE’s resistance to biodegradation, especially in
comparison to the BTEX components of gasoline. MTBE either did not biodegrade within the time limit of
within
the study, or it was biodegraded poorly, or under specific conditions, generally not found in natural
groundwater sediments. For example, some studies (Jensen and Arvin, 1990; Salanitro et al., 1994; Cowan
(Jensen

and Park 1996; Mo et al., 1997; Hanson et al., 1999; Pruden et al., 2001; Francois et al., 2002; Zaitsev et
al., 2007) have been conducted with microbial cultures derived from activated sludge (material, or flocs,
containing high concentrations of diverse groups of organisms, mostly bacteria, but also fungi and
mostly
protozoa, produced during the secondary treatment process of wastewater, i.e. sewage treatment). In other
words, the results of such studies cannot be applied directly to groundwater environments that lack the

population numbers and activity of microorganisms commonly found in activated sludge. Moreover, and as
sludge.
will be explained in the following sections, many studies which showed evidence of MTBE biodegradation
were conducted with surface water sediments (lakes, rivers, estuaries, etc…) and not aquifer sediments.
estuaries,
After nearly two decades of research on MTBE biodegradation, the consensus remains that MTBE is not an

easily biodegradable compound (McKelvie et al., 2007; Youngster et al., 2008; Waul et al., 2008;
Malandain et al., 2010), and convincing proof of degradation under natural groundwater conditions
(without engineered intervention) is scarce. Research is ongoing into the mechanisms involved in microbial
breakdown of the compound, enzyme induction and regulation, reasons for long acclimation periods when

54

biodegradation does occur, slow growth rates of microorganisms, as well as the nature and prevalence of
organisms capable of biodegradation, among other issues.
The following is a detailed discussion of the available literature on the biodegradability of MTBE. It should
be noted that in the presence of naturally occurring oxygen in groundwater, other gasoline constituents of
concern, particularly benzene, toluene, ethyl benzene and the xylenes (BTEX) would be biodegraded
preferentially (compared to MTBE), resulting in depletion of oxygen (USEPA, 2005). Consequently, “most
MTBE plumes are anaerobic” (USEPA, 2005). As such, in considering the biodegradation potential of
MTBE in aquifers, under natural non-engineered conditions, anaerobic conditions are of particular
significance for the natural attenuation of MTBE. And while dissolved oxygen concentrations as low as 0.1
mg/l may be enough for aerobic BTEX biodegradation, aerobic MTBE biodegradation appears inhibited at
concentrations less than 1.0 mg/l (Wilson, 2003) or 2 mg/l (Wei and Finneran 2009). Studies conducted
under aerobic conditions therefore, are necessary for understanding the potential for engineered
remediation of MTBE plumes where high dissolved oxygen concentrations can be maintained.

5.2.1 Anaerobic Biodegradation of MTBE
In the absence of oxygen, bacteria may use other naturally occurring chemicals to carry out biodegradation
reactions (anaerobic biodegradation). Such chemicals, referred to as electron acceptors, include nitrate
(NO3-), Manganese (Mn(IV)), Iron (Fe (III)), sulfate (SO4-), and carbon dioxide (CO2). Because the
“electron acceptor” becomes “reduced” as part of the reaction, the conditions under which such reactions
occur are referred to as reducing, e.g. nitrate reducing, sulfate reducing, iron reducing, etc…. When carbon
dioxide is used as the electron acceptor, methane is produced, and the reaction conditions are generally
referred to as methanogenic.
The different electron acceptors are used preferentially by bacteria. For example, for the BTEX group,
USEPA (2005) explains:
“Bacterial communities acclimate readily to degrade BTEX compounds using oxygen and nitrate
as electron acceptors. Depletion of oxygen and nitrate should be expected at almost every gasoline
spill. Bacterial communities also acclimate readily to degrade sulfate. Depletion of sulfate should
be expected at most sites as well. Bacterial communities require from months to years to acclimate
to ferment BTEX compounds to methane”.
In comparison, a close look at MTBE anaerobic biodegradation studies reveals contradictory results, as
well as the fact that there is little evidence to indicate that such processes have been documented in aquifer
sediments. Some of the studies that did show evidence of biodegradation under anaerobic conditions have
been conducted with surface water sediments, and not aquifer sediments, as discussed later. Furthermore,
biodegradation was not always established, even after very long acclimation periods.
Summary of Anaerobic Biodegradation Studies: One of the earlier studies (Mormile et al. 1994)
reported no biodegradation of MTBE in aquifer sediments, under nitrate reducing conditions, in 85 days of
incubation, and no biodegradation under sulfate reducing conditions in 244 days of incubation. Similarly,
55

Borden et al. (1997) reported no MTBE biodegradation in aquifer sediments in 263 days incubation under
low oxygen, nitrate rich conditions. Suflita and Mormile (1993) showed no biodegradation in aquifer
sediments under anaerobic/methanogenic conditions in 250 days of incubation. Similarly, Church et al.
(1999) found no evidence of MTBE biodegradation under anaerobic conditions in aquifer sediments from 4
different sites in 120 days. Finneran and Lovley, (2001) reported no biodegradation in aquifer sediments in
275 days, under sulfate reducing conditions when incubated without amendments (natural conditions). The
aquifer sediments degraded MTBE only after the addition of iron and humic substances, leading the authors
to conclude: “the aquifer sediments evaluated in this study also had little intrinsic potential to degrade
MTBE” (Finneran and Lovley, 2001).
On the other hand, research conducted by USGS scientists (Bradley et al., 2001a; Bradley et al., 2001b) on
surface sediments from different sources (lakes and streams) showed varying degrees of biodegradation of
MTBE, depending on source, incubation and redox conditions, under nitrate reducing, sulfate reducing,
iron reducing, manganese reducing, and methanogenic conditions. Meanwhile, Somsamak et al. (2001)
showed biodegradation of MTBE under sulfate reducing conditions in one out of two microcosms, with
estuarine sediments, after more than 3 years of incubation. However, in the same study, the sediments taken
from Arthur Kill estuarine inlet (located between Staten Island, New York and New Jersey) showed no
evidence of MTBE biodegradation under nitrate reducing, iron reducing, or methanogenic conditions in
three years of incubation. In another study, Finneran and Lovely (2001) showed biodegradation of MTBE
under iron reducing conditions in sediments from the Potomac River after an acclimation period of 250 to
300 days. Somsamak et al. (2005) showed biodegradation of MTBE in only 1 out of 3 microcosms with
estuarine sediments under anaerobic/methanogenic conditions, after an acclimation period of 390 days.
In a recent study (Waul et al., 2008), experiments were conducted on the anaerobic biodegradation of
MTBE both in batch, and continuous bioreactor settings, and under various electron acceptor conditions,
including nitrate, sulfate and iron (Fe(III)) reducing conditions, as well as methanogenic conditions.
Several sources of inocula, presumably rich in anaerobic bacteria, including sludge and contaminated soil
form a petroleum refinery that handles MTBE, were used “such that the best conditions existed for the
detection of anaerobic MTBE degradation.” The batch experiments were incubated for periods ranging
from 2 months to 2.5 years, whereas the bioreactors were operated for periods of 2 months and 3.5 years.
The researchers (Waul et al., 2008) reported no MTBE biodegradation in the bioreactors under any of the
redox conditions specified. In the batch experiments meanwhile, no degradation was reported under any of
the redox conditions, except for the Fe(III) reducing conditions where a 30 to 60% conversion of MTBE to
TBA was observed. Those batch experiments were not buffered and Waul et al., (2008) concluded that:
“The use of complexed Fe(III) created an initial low pH of 1-2 in these batches due to its acidic nature,
therefore the removal may be due to acid hydrolysis rather than biological processes.” The degradation of
MTBE to TBA through hydrolysis under acidic conditions has been reported previously (O’Reilly et al.,
2001; Zwank et al., 2005; Fischer et al., 2007).

56

Wei and Finneran (2009) were the first to establish a stable liquid anaerobic culture derived from
sediments from petroleum-contaminated groundwater from a gasoline station site in California.
Enrichments were prepared by the addition of the electron acceptors anthraquinone-2,6-disulfonate
(AQDS), fumarate, nitrate and sulfate. The samples enriched with either AQDS, fumarate, or sulfate
degraded MTBE after “significant lag periods for each of the electron acceptors tested, ranging from 240 to
315 days prior to complete degradation”. In each case, “TBA was produced as a result of MTBE
degradation, but was not itself degraded under any of the conditions tested”. Interestingly, Wei and
Finneran (2009) reported no MTBE biodegradation in the “nitrate-amended enrichments”. Furthermore, it
appeared that the added amendments (AQDS, fumarate or sulfate) were necessary for the cultures to grow
on MTBE; incubations without amendments were not able to grow on MTBE.
The researchers (Wei and Finneran 2009) tried to identify individual pure anaerobic microorganisms
responsible for MTBE biodegradation, but were not able to. Analysis of the microbial community
“indicated that the dominant populations in the MTBE-degrading cultures are not closely related to known
species, suggesting that anaerobic microorganisms growing on MTBE are in fact novel” [emphasis as
published]. Similarly, Youngster et al., (2010) tried unsuccessfully to isolate pure cultures degrading
MTBE, enriched from surface water sediments, noting that: “[c]haracterization of the MTBE-utilizing
enrichment cultures was hampered by their very slow growth…..the low population density contributes to
the difficulty in sustaining anaerobic MTBE-degrading activity in cultures over successive transfers and
dilutions.” The authors (Youngster et al., 2010) also speculated that interactions between more than one
species of organisms “may be required for appreciable anaerobic MTBE degradation to occur.”
It should be noted that from a theoretical standpoint, the oxidation of MTBE under the different electron
accepting conditions is thermodynamically feasible (Schmidt et al., 2004). However, the thermodynamic
feasibility of a reaction does not account for kinetic limitations and the ability of microorganisms to break
down and metabolize organic chemicals, or even the presence of degrading microorganisms. As evidenced
by the results of years of research so far on MTBE biodegradation, the presence of electron acceptors is not
necessarily enough for biodegradation to proceed.
In summarizing findings from research studies on MTBE biodegradation, USEPA (2005) cautioned:
“Although these studies prove that anaerobic biodegradation of MTBE in sediments is possible, they do not
indicate that anaerobic biodegradation in aquifer sediments is common or pervasive.” Similarly, Somsamak
et al., (2005) cautioned: “Considering that anaerobic MTBE degradation is very rare, the enzyme system
involved in attack may not be widely distributed”. Stupp (2007) also concluded “[a]ccording to field
observations in groundwater, MTBE is hardly microbiologically degradable by pure natural processes
alone.” Hence it would be imprudent to entrust such poorly understood activity with plume attenuation at
appreciable rates.
Biodegradation in Aquifer Sediments Compared to Surface Water Sediments: The results of MTBE
degradation in surface water sediments, as compared to aquifer sediments, need to be considered in the
57

context of the different environments (for example, microbial abundance, biodiversity), and how that
influences microbial activity. For example, a study by Kazumi and Capone (1994) on microbial population
and activity in the Upper Glacial Aquifer on Long Island (using glucose, a simple sugar), concluded:
“bacterial activity and abundance estimates [in the Upper Glacial Aquifer] are similar to values found in
other aquifer environments, but are 10- to 1000-fold lower than values in soil or surface sediment of marine
and estuarine systems” (Kazumi and Capone 1994). Consequently, for a chemical that is more difficult to
biodegrade (e.g. MTBE as compared to glucose), the disparity in response between the two environments
(surface versus aquifer sediments) would be expected to be more pronounced because of the nature, and
numbers of organisms available in each environment.
In fact, a review by Kieft and Phelps (1997) noted that: “microbiological activities in uncontaminated
subsurface habitats proceed at rates that are orders of magnitude slower than those encountered in nearly
every other portion of the biosphere”. Based on the work of Kieft and Phelps (1997), a more recent review
of “Microbial Biodiversity in Groundwater Ecosystems” by Griebler and Lueders (2008) noted that: “in
general, differences in microbial activities between subsurface and surface environments can reach up to 10
orders of magnitude (Kieft and Phelps, 1997)”. Griebler and Lueders (2008) also noted that published
reports “support the idea that indigenous groundwater microbial communities are distinct from those found
in surface environments.” As such, results of biodegradation experiments conducted in the laboratory with
samples from surface water, or sludge from wastewater treatment plants, cannot be directly extrapolated to
subsurface environments without further evidence from aquifer studies, particularly when it is an
anthropogenic chemical like MTBE, that is not easily degradable.

5.2.2 Abiotic degradation and apparent biodegradation under methanogenic
conditions
There are some reports of anaerobic degradation of MTBE in aquifer sediments that were conducted under
methanogenic conditions in what is referred to as an “anaerobic glove box” (Wilson et al., 2000; Kolhatkar
et al., 2002; Wilson et al., 2005b). An “anaerobic glove box” refers to incubation under an “anaerobic
mixed gas” consisting of carbon dioxide, nitrogen, and up to 5% by volume Hydrogen (Bradley et al.,
2006). Recently, USGS scientists (Bradley, Landmeyer, and Chapelle in Bradley et al., 2006), the same
group that conducted the surface water sediment experiments showing MTBE biodegradation under
different electron accepting conditions (Bradley et al., 2001a; Bradley et al., 2001b; Bradley et al., 2001c),
showed that under conditions similar to those enforced in an anaerobic glove box, specifically relatively
high hydrogen concentrations, which is also typical of methanogenic aquifer conditions, the transformation
of MTBE to TBA is likely chemically (abiotic), and not microbially mediated. Interestingly, in their own
work previously (Bradley et al., 1999) with surface sediments under methanogenic conditions, incubated
with Helium gas instead of in an anaerobic glove box, no biodegradation of MTBE was observed in
sediments from 2 different sites in 105 days of incubation.
In the field, fermentative microorganisms produce hydrogen as they decompose organic matter under

58

anaerobic conditions (Chapelle et al., 1996). Methanogenic conditions are characterized by hydrogen
concentrations in the range of 5 to 15 nM (Chapelle et al., 1996). In their work, Bradley et al. (2006)
showed that abiotic transformation of MTBE to TBA occurs at hydrogen concentrations above 5nM. The
experiments included comparisons of MTBE transformation in “live” (non-autoclaved) microcosms to
microcosms repeatedly autoclaved (heat sterilized) to kill or suppress microbial activity. The autoclaved
samples included microcosms with hydrogen (not flushed), and microcosms flushed with nitrogen to
remove the hydrogen (flushed). The results showed that MTBE transformation in the “live” microcosms
was almost identical to that in the autoclaved microcosms not flushed (containing hydrogen). In
comparison, the samples that were autoclaved, but flushed with nitrogen to eliminate hydrogen, showed no
transformation. In other words, in the absence of hydrogen, there was no transformation of MTBE, leading
Bradley et al. (2006) to conclude that at hydrogen concentrations greater than 5nM (typical of
methanogenic conditions), abiotic reduction, and not biological degradation, of MTBE to TBA is likely the
primary transformation pathway.
The results by Bradley et al. (2006) suggest that under natural aquifer conditions, biologically produced
hydrogen can facilitate MTBE reduction to TBA and methane, without the reaction being microbially
mediated. This is also particularly important since it is common practice to add hydrogen gas to the
headspace of laboratory microcosms incubated under anaerobic conditions (Bradley et al., 2006). Given
that the anaerobic biodegradation of MTBE has yet to be connected to specific microorganisms, or clear
enzymatic pathways, this puts into question published literature on anaerobic MTBE biodegradation under
methanogenic conditions. It produces doubt on whether MTBE transformation to TBA under high
hydrogen concentration conditions is mediated by microorganisms or is simply a chemical reaction
catalyzed by other elements within the sediment matrix. Importantly, under such conditions TBA is likely
to be a dead end product. TBA biodegradation under anaerobic conditions has not been well documented in
field studies (USEPA, 2005).

5.2.3 Aerobic biodegradation
In one of the earliest laboratory studies, Jensen and Arvin (1990) used 4 sources of inocula rich in
microbial cultures (soil, sandy aquifer material, and two types of activated sludge), to examine the
biodegradability of typical gasoline aromatics (BTEX and naphthalene) as well as MTBE. The aromatics
were completely degraded within 3 to 13 days, but no MTBE biodegradation was observed in 60 days.
Probably the first study to demonstrate successful mineralization (transformation to carbon dioxide and
water) of MTBE under laboratory conditions was published by Salanitro et al. (1994). In this study, three
types of activated sludge (municipal, refinery, and chemical plant bioreactors) were used in multiple
inoculations to enrich for a microbial culture that degrades MTBE. After 2 months, the chemical plant
activated sludge developed an MTBE-degrading culture. The bacterial culture however was found to have a
low growth rate and a low cell yield. Cowan and Park (1996) also developed a mixed microbial culture that
degraded MTBE using inoculum from activated sludge from a petroleum-refinery wastewater treatment
59

plant. Rates of MTBE biodegradation were higher at 30oC (86oF) compared to 20oC (68oF) (Park and
Cowan, 1997). The rate of biodegradation was also found to be dependent on oxygen concentrations, with
rates dropping significantly at dissolved oxygen concentrations below 2.1 milligrams per liter (mg/l).
Several other studies (Borden et al., 1997; Bradley et al., 1999; Salanitro et al. 2000; Feris et al., 2004;
Chen et al., 2007) have also noted the process of biodegradation was dependent on oxygen.
However, in many other cases, MTBE was reported to be resistant to biodegradation in the presence of
oxygen. For example, in a controlled laboratory experiment, Pasteris et al. (2002) noted that out of a
mixture of 13 fuel compounds, MTBE was the only one to persist in groundwater through 70 days of
monitoring. In another study (Ruiz-Aguilar et al., 2002), no evidence of MTBE biodegradation was
observed in aquifer material collected from 4 different sites, during 100 days of incubation (both under
aerobic and anaerobic conditions). Da Silva and Alvarez (2002) also reported no biodegradation of MTBE
in biologically active columns in a period of 110 days. Similarly, Dakhel et al. (2003) reported that MTBE
persisted for longer than 6 months with no evidence of biodegradation.

5.2.4 Engineered Bioremediation
Given that “most MTBE plumes are anaerobic” (USEPA, 2005), and that the biodegradation of MTBE is
more likely to proceed at appreciable rates under aerobic conditions, several field studies have been
conducted on the efficacy of “oxygenation” as a remediation option (for example Salanitro et al., 2000;
Wilson et al., 2002; Landmeyer et al., 2003; and Smith et al., 2005). In this case, several site-specific
conditions will influence process performance. For example, Landmeyer and Bradley (2003) showed that
the addition of oxygen release compounds did not result in changes in the dissolved oxygen or MTBE
concentrations in the source zone, but did result in measurable reduction in MTBE concentrations 200 m
down-gradient from the source of contamination. This difference in response at the two locations was
attributed to difference in the hydrologic, geochemical and microbiological characteristics of the two zones
of contamination, including redox conditions, and the nature of the microorganisms present at the two
locations.
At the contaminated site in Port Hueneme, California, Salanitro et al. (2000) conducted a pilot test to study
the effect of oxygen injection alone, as compared to oxygen injection, combined with bioaugmentation
(addition of proprietary MTBE degrading microorganisms) on the removal of MTBE in situ. Results
indicated that the bioaugmented plot performed better (lag time approximately 30 days), compared to the
oxygen only plot (lag time 186 to 261 days). A follow-up study published recently (Lesser et al., 2010)
compared the activity of MTBE degradation in different cores from the site. The authors (Lesser et al.,
2010) noted that: “cores from the bioaugmented regions had the highest frequency of MTBE
biodegradation to nondetect levels among their sections suggesting a direct effect of the inoculum and its
distribution when it was implanted”. Smith et al. (2005) conducted a similar experiment at Port Hueneme.
The addition of oxygen alone in one plot was compared to the addition of oxygen combined with
bioaugmentation with the UC Davis MTBE degrading bacterial strain PM1. Similar results were observed
60

with less lag time in the bioaugmented plot. In both studies oxygen delivery was problematic at depths
greater than approximately 15 feet, limiting its applicability to shallow contamination. At the same time,
the large network of wells necessary to create a “biobarrier” can limit the applicability of such a process
because of the need for site access for installation, operation and maintenance. Other site-specific
conditions would have to be considered. For example, heterogeneity and low permeability sediments, as
well as large fluctuations in groundwater levels may complicate treatment. As Wilson et al. (2002)
concluded: “We believe a wide range of methods are available to create aerobic conditions [in situ]; what is
needed is an assessment of the efficacy of each method for a range of hydrogeologic and geochemical
conditions and continued work to understand the prevalence and reliability of native MTBE-degrading
microbes.”
When the contaminated water is deep underground, ex-situ treatment may be necessary. Ex-situ treatment
involves pumping groundwater, and treating it above ground. More recently Tornatore et al., (2005; 2006)
reported on the use of granular activated carbon (GAC) beds to seed native MTBE degraders. By providing
oxygen and nutrients, bacteria in the groundwater, capable of MTBE biodegradation, created a biofilm that
is capable of treating MTBE and TBA. This project, in Los Angeles County, is particularly unique because
the treated water is allowed to be re-injected in the subsurface, rather than be disposed of off-site. This
approach, possible because the degrading microorganisms are native to the aquifer, serves to preserve the
water source. Re-injection of oxygenated water may also stimulate in-situ biodegradation.

5.3

Factors Affecting Biodegradation

Based on the scientific literature, several factors appear to be necessary for MTBE biodegradation to occur
in the subsurface, assuming the presence of native MTBE degraders. These factors include pre-acclimation
to MTBE, the presence of sufficient oxygen, as well as other site-specific conditions, such as redox
conditions, pH, and temperature.

5.3.1 Difficulty obtaining a degrading microbial culture: While evidence from
some sites indicates that natural MTBE degraders are present and can be stimulated with the addition of
oxygen (Salanitro et al 2000; Wilson et al., 2002; Smith et al., 2005), many researchers have noted that
MTBE degrading microorganisms are not prevalent even at contaminated sites exposed to MTBE for long
periods of times. “Generally it has proved difficult to isolate strains from enrichment cultures using MTBE
as a sole carbon and energy source. Attempts for more than 15 years were of limited success” (Muller et
al., 2008).
Unlike BTEX degrading microorganisms, microorganisms that biodegrade MTBE aerobically do not
appear to be prevalent at contaminated sites. “Fuel oxygenates such as methyl and ethyl tert-butyl ether
(MTBE and ETBE, respectively) are degraded only by a limited number of bacterial strains” (Rohwerder et
al., 2006). Many of the aerobic microorganisms that have been identified as MTBE degraders were
cultured from activated sludge (from sewage treatment plants) (Salanitro et al., 1994; Cowan and Park

61

1996; Mo et al., 1997; Hanson et al., 1999; Pruden et al., 2001; Francois et al., 2002; Zaitsev et al., 2007).
A study by Pruden et al. (2001) indicated that the group of organisms identified in MTBE biodegradation
may be dominant in activated sludge, but may not be present in uncontaminated aquifers. Furthermore,
microorganisms that degrade MTBE anaerobically have not been isolated. Youngster et al., (2010) noted
that: “[t]o date, very little is known about the anaerobic MTBE degradation process and none of the
responsible organisms have been identified or isolated as yet”.
Spinnler et al. (2001a), for example, reported that no MTBE degrading microorganisms were found in
samples collected from 2 contaminated sites, one in the Northeast, and one in Northern California. Spinnler
et al. (2001b) stated: “Microorganisms capable of degrading MTBE are not so ubiquitous as BTEX
degraders. … few soils collected from MTBE contaminated sites showed any MTBE-degrading activity
when stimulated with oxygen.” Salanitro et al. (1998) made similar observations. They found that 8 out of
10 soil and groundwater samples collected from different contaminated sites from around the country, had
no MTBE-degrading microorganisms. Activated sludge from 4 refineries (out of 5 tested), also had no
MTBE degrading capability. Salanitro et al. (1998) concluded: “Our survey results … on the presence of
ether-degraders in soil and ground water show that these populations are too low and/or grow too slowly to
effectively attenuate observed levels of MTBE in aquifer plumes.” Lesser et al., (2010) noted differences
in degradation activity at the Port Hueneme site within small distances: “significant changes in activity
level were observed over 0.3-m [~ 1 foot] vertical distances in the same location.”
Kane et al (2001) examined aquifer sediments from 7 MTBE contaminated LUFT sites in Northern
California, noting that only 3 out of 7 showed evidence of MTBE biodegradation. Similarly, Moreels et al.
(2004) examined the potential for MTBE biodegradation, under aerobic conditions, in soils collected from
seven different locations in Belgium. Five of the samples were collected from sites contaminated with
gasoline, while two were topsoil samples. MTBE was degraded in samples from only one of the sites, with
a lag-phase of up to 270 days. In contrast, benzene was degraded rapidly in all samples from all sites.
More recently, Babe et al., (2007) examined the capacity of native microorganisms to biodegrade MTBE
and other oxygenates in soil and groundwater samples from 14 different locations in Europe and the United
States. All samples came from “fuel oxygenate contaminated sites”. Optimum growth conditions were
provided with samples incubated under aerobic conditions (with oxygen in “large excess” of the theoretical
demand) at 30oC, and with yeast extract added as a supplement to improve microbial growth. Results
showed “either no or very low capacities for MTBE biodegradation” in nine out of the fourteen (64%) site
samples. Out of the samples collected from the USA (seven out of the fourteen) only 2 sites showed
efficient biodegradation. Three sites showed no biodegradation, while two others showed a low capacity to
degrade MTBE (9% MTBE degraded in 345 days in one soil sample, and 21% MTBE biodegraded in 277
days in an aquifer sample from another location) Babe et al., (2007).
Based on observations from the published literature on the difficulty in establishing MTBE biodegradation
under natural anaerobic aquifer conditions, the difficulty isolating anaerobic microorganisms degrading
62

MTBE, the apparent contradictory results from surface sediments (some positive others negative), and the
relative abundance of studies with degrading microorganisms isolated from activated sludge, combined
with the apparent absence of aerobic MTBE degrading organisms from many contaminated sites it may be
these observations can be explained partly by the specificity of organisms that can degrade MTBE. In his
book “Biodegradation and Bioremediation” Alexander (1994) explains:
“It also is now clear that microorganisms acting on certain pollutants are absent from particular
sites. A compound that is metabolized by many species will likely encounter one or several
species in all microbial communities that can transform it. On the other hand, certain synthetic
compounds are apparently transformed by very few species, and it is thus likely that not a single
one of the very few species with the requisite enzymes may be present in a particular site. This
view is in line with the frequent observation that some organic compounds are mineralized or
otherwise metabolized in samples from one but not another environment and that active organisms
can only be isolated from some environments.”

5.3.2 Acclimation: Also referred to as adaptation, or lag time, is the time needed for
microorganisms to display measurable biodegradation. Acclimation is important in understanding when
and under what conditions biodegradation will proceed. Becker et al., (2006) explain: “current
understanding of the events that lead to microbial community adaptation to pollutants is extremely limited.
This lack of understanding means that we are unable to predict when or where adaptive events will occur.”
Under ideal (e.g., controlled laboratory/oxygenated) conditions, time lags for MTBE biodegradation of 200
days (Mosteller et al., 1997), 260 days (Salanitro et al., 2000), and up to one year (Eweis et al., 1997) have
been reported. Such observations suggest that the lag times in natural aquifers, typically under oxygen and
nutrient limited conditions, will likely be even longer, maybe years (USEPA, 2005). The reasons for a long
acclimation period remain unclear, but even assuming MTBE biodegradation does occur, a lengthy time lag
will allow the leading edge of MTBE contamination to advance without any attenuation by biodegradation.
From experience with a field experiment on MTBE biodegradation in the Borden Aquifer in Canada,
Schirmer et al. (2003) concluded:
“It is possible that the lag periods reflect slow growth of a population originating from a very
small number of cells, so that much time elapses before detectable MTBE loss. An alternate
explanation of these findings is that fortuitous genetic change (perhaps mutation) plus selective
pressure (MTBE as the only available organic substrate) has led to an MTBE-degrading capacity.
If so, though useful when or if it occurs, this phenomenon would be impossible to predict. Natural
attenuation of a MTBE plume as a result of such occurrences could never be relied on.”

5.3.3 Oxygen: In some cases, low oxygen concentrations have been associated with reduced or
inhibited MTBE biodegradation (Scow et al., 2000; Landmeyer et al., 2001), and the addition of oxygen
has been shown to stimulate MTBE biodegradation (Salanitro et al., 2000 and Wilson et al., 2002). At a
site in South Carolina (discussed earlier in this document), Landmeyer and Bradley (2003) found that the
addition of oxygen release compounds (to increase dissolved oxygen concentration in groundwater) did not
result in changes in the dissolved oxygen or MTBE concentrations in the source zone. It did, however,

63

result in a measurable reduction in MTBE concentrations 200 m down-gradient from the source of
contamination. This result (difference in response) at the two locations (source zone versus down-gradient)
was attributed to differences in the hydrologic, geochemical and microbiological characteristics of the two
zones of contamination.
Simply put, the presence, or introduction of oxygen, is not necessarily enough to induce MTBE
biodegradation. In many cases, MTBE was reported to be persistent even in the presence of oxygen
(Pasteris et al., 2002; Da Silva and Alvarez 2002; Dakhel et al. 2003; Moreels et al., 2004; Babe et al,
2007). In fact, a recent analysis (Squillace et al., 2007 and Carter et al., 2008) of data from 3,500 water
samples collected between 1985 and 2002 from almost 100 different aquifer studies over the U.S., shows
that MTBE detections in groundwater seem to be more highly associated with oxic groundwater, as
compared to anoxic groundwater. Oxic groundwater was defined as having a dissolved oxygen (DO)
concentration greater than 0.5 mg/L, while anoxic groundwater had a DO of less than or equal to 0.5 mg/L
(Squillace et al., 2007). It should be noted that: “Young ground water usually has a larger dissolved
oxygen concentration compared to old ground water” (Zogorski et al., 2006, emphasis as published).
Young ground water was defined as “water recharged after 1955”. In other words, the association between
MTBE detections and groundwater type (i.e., oxic versus anoxic) could have more to do with the relatively
recent introduction of MTBE into the environment (young groundwater), and less to do with its
biodegradation potential.

5.3.4 Other Factors: In some cases researchers have found that MTBE biodegradation is
inhibited in the presence of more easily degradable compounds, such as the BTEX group (Church et al.,
2000; Deeb et al., 2001; Auffret et al., 2009), or alkane compounds (Johnson and Hyman, 2006). The pH
may also play a role; Pruden et al. (2001) reported an optimum pH range of 7.4 to 7.6, with rapid
deterioration of MTBE removal efficiency outside of the pH range of 7.3 to 8.0. Muller et al., (2008) on the
other hand, reported growth of a pure MTBE degrading strain within a pH range of 5.5 to 8.0, but no
growth at pH below 5.0, and above 9.0. Eweis et al. (1997) reported similar results. At the same time,
temperature can be a factor in biodegradation rates. For example, while Park and Cowan (1997) reported
higher rates of MTBE biodegradation at 30°C compared to 20°C (in laboratory microcosms under aerobic
conditions), Bradley and Landmeyer (2006) reported optimum biodegradation at 24°C with significant
decrease in biodegradation potential at temperatures below, or above 24°C, under similar laboratory
conditions. At the lower temperatures (as low as 4°C) biodegradation was significantly reduced, but not
completely inhibited (Bradley and Landmeyer, 2006).
Another important factor is what appears to be the slow growth rates associated with MTBE degrading
organisms. Muller et al., (2007) suggested that a relatively high threshold concentration might be necessary
to drive growth on MTBE. For the most part, organisms that have been shown to grow on MTBE are
relatively slow growing, with low cell yield (Salanitro et al., 1994; Mo et al., 1997; Hanson et al., 1999;
Fortin and Deshusses, 1999; Pruden et al., 2001; Zaitsev et al., 2007; Vosahlikova-Kolarova et al., 2008;
64

Youngster et al., 2010). Rates of biodegradation are inherently tied to growth rates. USEPA (2005)
explains:
“Typical strains of bacteria growing aerobically on petroleum hydrocarbons can divide and double
their numbers every two to five hours at room temperature. As a consequence, laboratory
enrichment cultures will grow up and remove the hydrocarbons in a few days. On the other hand,
cultures of bacteria using MTBE as a growth substrate require several days to several weeks to
double their numbers. Their growth rate is from one-tenth to one-hundredth of the growth rate of
bacteria that degrade conventional petroleum hydrocarbons. Their very slow growth rate has an
important effect on the time required for a culture to grow to densities that will entirely consume
MTBE.”
Some studies have suggested that some amendments may be necessary to support MTBE biodegradation;
for example cobalt (Rohwerder et al., 2006), or humic substances (Wei and Finneran, 2009). Other factors
influencing MTBE biodegradation remain unclear. For example, in some laboratory studies, the
biodegradation of MTBE was observed to cease, after it had progressed for some time, and at relatively
high concentrations (Mo et al., 1997; Borden et al., 1997; Schirmer et al., 1999). In those studies,
researchers could not identify a clear cause for the cessation of microbial activity.

5.4

Examples from field studies

Evidence of MTBE biodegradation in aquifers under natural conditions (in the absence of engineered
remediation) remains scarce. Below is a summary of a few examples from field studies that are of
particular interest because of their history or site-specific conditions that might influence MTBE fate.
Canadian Forces Base, Borden Ontario: A field experiment at this site in 1988, resulted in the controlled
release of gasoline containing MTBE in the shallow aquifer. Monitoring over a period of 476 days showed
MTBE to be recalcitrant, moving at the same rate as a conservative tracer (Barker et al., 1990). The
researchers then re-sampled groundwater from the site in 1995-96, nearly 7 years from the time the plume
had last been sampled (Schirmer and Baker, 1998). With limited sampling on a grid much coarser than that
of the original experiment, they could not account for all of the MTBE mass that was originally injected.
They indicated that this apparent mass loss may have been the result of biodegradation. However, because
of the large time span between the sampling events and the lack of documentation of other processes that
could be responsible for mass disappearance, the case for biodegradation remains weak. Other processes,
such as diffusion into the underlying aquitard (which the plume had migrated close to), transient
fluctuations of the hydraulic gradient causing plume migration into unanticipated locations and greater
mixing, and relatively extreme precipitation events could have affected plume movement and
concentrations. In fact Schirmer and Barker (1998) stated:
“While we are confident that MTBE mass has been lost, we cannot yet confirm biodegradation as
the process.” In reference to the same study site, Schmidt et al. (2004) remarked: “even at a site
with injection of a defined amount of the gasoline oxygenate methyl tert-butyl ether (MTBE) into
groundwater, well-known hydrogeology, and a dense network of sampling wells, it was not
possible to unequivocally determine MTBE biodegradation after a period of eight years”.
65

Former UST area near Beaufort, South Carolina: The importance of oxygen in MTBE biodegradation
was further emphasized with several studies conducted at a site in South Carolina (Landmeyer et al., 1997,
Landmeyer et al., 1998, and Landmeyer et al., 2001) where a plume of MTBE extends approximately 230
m (755 feet) down-gradient from a UST area, to where groundwater discharges into a concrete-lined ditch
(surface water). Gasoline was detected in the shallow aquifer in 1991, “although the UST had most likely
leaked since the mid-1980s” (Landmeyer et al., 1998). The study published in 1998, approximately 10
years after the release, examined the biodegradation of MTBE in microcosms containing sediments from
the contaminated aquifer. The experiments were conducted using radio labeled MTBE. In a 7-month
period, less than 3% of the radiolabeled MTBE was recovered as radiolabeled CO2 (evidence of limited
biological transformation to CO2). Biodegradation was deemed insignificant as a natural attenuation
mechanism in the contaminated aquifer. In a later study (Landmeyer et al., 2001), MTBE biodegradation
was documented in samples from a microbial biofilm growing in the concrete-lined ditch (zone exposed to
oxygen) where groundwater discharges into surface water. In other words, the conditions in the aquifer
itself did not favor natural biodegradation, and biodegradation only occurred when water became
oxygenated at the discharge surface.
Deer Park, New York: At this site, the release of gasoline was first detected in 1990, at a time when
MTBE was used as an octane enhancer at lower concentrations. Thuma et al. (2001) reported that the
BTEX plume extended approximately 700 feet, whereas the MTBE plume extended to about 8000 ft. Initial
remediation at the site included excavation of the impacted soil and dual phase extraction (Wilson and
Kolhatkar 2002). Based on the 2002 report the site was continuing to be remediated with air sparging and
soil vapor extraction.
Kolhatkar et al. (2001) looked at MTBE concentrations in monitoring wells in the methanogenic zone of
the plume, and “calculated” an MTBE attenuation rate, from which a rate of apparent biodegradation was
further estimated. Kolhatkar et al. (2001) concluded: “[T]he ground water data in this flow path seem to
suggest that MTBE is being naturally biodegraded in the methanogenic ground water. Laboratory
microcosm experiments have been set up to further evaluate this possibility” However, the results of these
laboratory experiments as reported later by USEPA (2005) were in fact negative.
Thuma et al. (2001) developed a model for the site to simulate fate and transport in the plume and
investigate future remedial options, including monitored natural attenuation. The model assumed a
biodegradation rate of 5.2/year (estimated value from Kolhatkar et al., 2001) in the “methanogenic” zone
of the plume. The model did not seem to incorporate the effects of remediation processes already on site
(air sparging and soil vapor extraction). Thuma et al. (2001) concluded: “The predicted model
concentrations generally correlate well with field data in the upgradient half of the plume, however, further
downgradient the model predicts higher concentrations than have been observed in the downgradient
monitoring wells. We think this may be indicative of a methanogenic biodegradation zone extending
further downgradient than assumed, or occurrence of MTBE biodegradation downgradient of the

66

methanogenic zone, but at a slower rate, or the importance of sorption retardation in attenuating MTBE”.
However, in a follow up analysis, Wilson and Kolhatkar (2002) attributed the observed attenuation at the
site, in the distal part of the plume, “almost entirely to the contribution of dispersion.” No tangible evidence
of biodegradation was established, and the model did not match the field data well. Moreover, laboratory
microcosms taken from the site showed no evidence of biodegradation (USEPA, 2005).
Hampton Bays, New York: The plume at Hampton Bays was discovered when 54,000 µg/L MTBE and
other gasoline chemicals were detected in a well installed by Suffolk County Department of Health
Services (SCDHS) as part of a pesticide monitoring program (EAR, 2005). An investigation of the site
revealed the source of contamination to be a gasoline retailer, approximately 1000 feet up-gradient, with
MTBE concentrations near the source of 450,000 µg/L (EAR, 2005). The MTBE plume discharged into
Tiana Bay approximately 2,200 feet from the source. “Concerns for potential impact to the benthic biota of
Tiana Bay, from the discharging contaminated groundwater, caused NYSDEC to undertake several interim
remedial measures (IRM’s).” (EAR, 2005). The remedial measures included several treatment techniques,
including Bio-GAC18 for above ground water treatment.
To assess the potential for biodegradation in the plume, samples were collected and analyzed for evidence
of isotopic fractionation. After three rounds of sampling, over the span of approximately one year, the
results of the isotopic fractionation study concluded that: “[T]he isotope data and contaminant
concentration data for the most recent sampling round indicated that dilution and dispersion are likely the
most dominant attenuation mechanisms for MTBE removal within the aquifer” (Finneran, 2004, in EAR,
2005). The isotopic signal for δ13C% MTBE, in three rounds of sampling, ranged between –30.9 and –29.1,
which is within the range of –27.5% and –33% reported for MTBE in gasoline (USEPA, 2005), indicating
no apparent enrichment. The lack of evidence of any significant biodegradation under either aerobic
(dissolved oxygen concentrations up to 4 mg/L at leading edge of the plume), or anaerobic conditions at
this site, were further confirmed in a recently published report (McKelvie et al., 2007): “biodegradation of
MTBE is not a significant process at this site under either aerobic or anaerobic conditions”. In a more
recent study (Landmeyer et al., 2010), the potential for biodegradation of MTBE (and TBA and TAME) in
the hyporheic zone at Tiana Bay (where the plume discharges) was examined. The hyporheic zone, as
described by Landmeyer et al., (2010), is “the highly variable zone” in surface water bodies “where surface
water flows through parts of the bed and bank material and thus can interact with discharging
groundwater”. The authors also described the hyporheic zone as “a location of increased biological
abundance and diversity”. Although laboratory controlled studies showed the potential for MTBE

18

A BIO-GAC differs from a conventional granular activated carbon (GAC) bed, which is designed to
remove chemicals from water through adsorption (physical removal), in that a Bio-GAC will allow
microorganisms to grow on the surfaces of the granules removing chemicals biologically, once the
adsorption capacity of the bed is reached. In one of the Bio-GAC systems 1,350 gallons of Bio-Remedy
biomass® (Shell Global Solutions patented MTBE degrading microorganisms), were added to the
activated carbon beds to help remove MTBE.
67

biodegradation in samples from the hyporheic zone19, the authors concluded that “the primary mechanisms
of natural attenuation observed in the hyporheic zone were dilution and dispersion”, with only 0.5% of
attenuation attributed to “apparent biodegradation” in Tiana Bay (Landmeyer et al., 2010). The study which
examined MTBE attenuation in the hyporheic zone at three other surface water bodies in Long Island, New
York, reached similar conclusions for the other sites stating: “the attenuation observed in the hyporheic
zone occurred primarily by physical processes such as mixing of groundwater and surface water” which is
“in contrast to immeasurable attenuation in the shallow aquifer during contaminant transport between 0.1
and 1.5 miles” (Landmeyer et al., 2010).
Leuna, Eastern Germany: At an old refinery site dating over 20 years, releases of MTBE-containing
gasoline resulted in widespread contamination, where the MTBE plume extends to about 1700 m (1.05
mile), and is several hundred meters wide. Martienssen et al. (2006) reported that MTBE biodegradation
appears to be occurring under “microaerobic conditions”. While BTEX and other hydrocarbons appear to
be degraded anaerobically, “there was no evidence for anaerobic MTBE degradation” (Martienssen et al.,
2006). Samples from the site could not degrade MTBE anaerobically (Waul et al., 2008) whereas pure
bacterial cultures isolated from the site were strictly aerobes (Lechner et al., 2007). Researchers
(Martienssen et al., 2006) postulated that conditions at the Leuna site, such as unconfined aquifer, close to
ground surface, result in oxygen replenishment following rainfall events and related groundwater rise,
leading to areas with elevated oxygen concentration, (microaerobic conditions) that favor aerobic
biodegradation. The researcher (Martienssen et al., 2006) concluded: “Since the oxygen supply is limited,
natural attenuation of MTBE requires large reaction zones.”
In a recent publication (Jechalke et al., 2010), researchers used contaminated groundwater “from a well
located downstream from the refinery” to examine the potential for biofilm formation and biodegradation
of benzene and MTBE in aerated ponds. Two different “geotextiles” were used as support material for the
biofilm in the aerated ponds: polypropylene fibers, and a natural coconut fiber. In the experiment that lasted
14 months, benzene was removed in both beds (from 20 mg/l in the inflow to about 1 µg/l in the outflow
with 1% of the removal being attributed to volatilization). In contrast, “MTBE biodegradation was by far
not as effective as benzene biodegradation” (Jechalke et al., 2010). Approximately 48-53% of MTBE
removal20 in the first 6 months was attributed to volatilization, although removal in the coconut fiber bed
increased to 67% after that, presumably due to biodegradation. Laboratory microcosm experiments were set
up using material from both types of beds (approximately 8 months after startup of pond experiments) to
examine/confirm biodegradation. Samples from the coconut fiber bed were able to degrade MTBE, though
“no biodegradation was observed for polypropylene after 100 days of incubation” (Jechalke et al., 2010). It
should be noted that despite the presence of “microaerobic conditions” at the Leuna site (as reported above
by Martienssen et al., 2006), Jechalke et al., (2010) noted that: “the inflowing groundwater is anoxic, and
19

Approximately 24% and 20% MTBE was degraded under aerobic and anaerobic conditions respectively,
in 83 days of laboratory incubation.

20

Based on an initial MTBE concentration of about 4mg/l.
68

major parts of the basins showed minimum oxygen concentrations between 0.0 and 1.0 mg/L.”
Meanwhile, at another site in Dusseldorf, Germany, Rosell et al. (2005) found no evidence of significant
natural attenuation of MTBE (no measurable mass loss) over a two-year monitoring period. TBA
concentrations at the site, at a level similar to that of MTBE, were attributed to its presence in the
contamination source (gasoline spill). TBA was found to be recalcitrant in the mostly anaerobic (dissolved
oxygen concentrations less than 1.1 mg/l) groundwater at the site.

5.5

Proxy Indicators of MTBE Biodegradation at Field Sites

Establishing clear evidence of MTBE biodegradation in the field is difficult. Microcosm studies which are
typically used to document the presence of degrading microorganisms in field samples, can be "expensive,
time consuming, and often yield equivocal results". (USEPA, 2005). Other evidence includes depletion of
electron acceptors such as oxygen, nitrate, sulfate, etc. However, because MTBE is usually present as part
of gasoline, the depletion of electron acceptors could be due to the consumption of BTEX (other gasoline
components), and not necessarily MTBE. Sometimes, the detection of a degradation product can offer
evidence of biodegradation. In the case of MTBE, TBA is "the primary degradation product of MTBE"
(USEPA, 2005). However, and as explained in the following section, TBA is a fuel oxygenate that might
have been added to gasoline intentionally. It is also a co-product of MTBE, and is often found as an
impurity in the MTBE added to gasoline.
Recently, "isotopic fractionation" has been employed as a tool to infer MTBE biodegradation at
contaminated sites. However, as discussed later, the application of this method to document MTBE
biodegradation at contaminated sites is a relatively new, and still evolving field of research. While
biodegradation in laboratory microcosms has been shown to result in isotopic fractionation, other physical
and chemical processes at play in contaminated sites, and that can also affect isotopic fractionation, had
been largely ignored until recently.

5.5.1 TBA at MTBE contaminated sites
At many sites, field data indicate that the frequency of occurrence of TBA and the range of concentrations
are at a scale similar to that of MTBE (Kolhatkar et al., 2000; Shih et al., 2004; Wilson et al., 2005a). For
example, Shih et al. (2004) found that for 868 randomly selected LUFT sites in the greater Los Angeles
area, MTBE was detected at 82.7% of sites, comparable to benzene at 82.5% of sites, while TBA was
detected at 61.1% of sites. At the sites studied, TBA was found to have the greatest site maximum
concentration, followed by MTBE, suggesting high source concentrations. TBA concentrations ranged
from 6 µg /l to 4,400,000 µg/l. On the other hand, MTBE concentrations ranged from 0.46 µg/l to
16,000,000 µg/l. As part of the study (Shih et al., 2004), 500 sites were considered for plume length
analysis, a measure of the potential for contaminants to affect drinking water sources. MTBE was found to
have the greatest plume length followed by TBA.
TBA is potentially more toxic than MTBE. In California, the notification level (previously known as the
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action level) for TBA is 12 µg/l based on evidence of potential carcinogenic properties (OEHHA, 1999).
Because of its properties (high mobility, low volatility, recalcitrance under natural field conditions) TBA
can be the regulatory driver for remediation considerations at contaminated sites (Shih et al., 2004), and is
expected to add to the cost of treatment of MTBE contaminated drinking water (AMWA, 2005; AWWA,
2005). Yet, testing for TBA at contaminated sites is a relatively new practice (Wilson et al., 2005a) and is
not required in most states. In 2003 only 11 out of the 50 states required sampling and analysis for TBA in
groundwater at gasoline contaminated LUFT sites (this compared to 31 states that required testing for
MTBE) (NEIWPCC, 2003). Hence, the level to which TBA has spread in groundwater at gasoline
contaminated sites remains largely unknown.
While TBA can be a product of MTBE degradation (Salanitro et al., 1994; Kane et al., 2001; Vainberg et
al., 2006; Skinner et al., 2008; Wei and Finneran 2009; Youngster et al., 2010), or reduction (Bradley et
al., 2006), as discussed earlier, it is also a co-product of MTBE and is found as an impurity in the MTBE
added to gasoline (Eweis et al., 2007). The TBA content in MTBE varies widely, and can be as high as
10% (Kramer and Douthit, 2000). Because TBA is completely miscible with water (no solubility limit),
even trace amounts of TBA in gasoline can result in large aqueous TBA concentrations, e.g., hundreds of
mg/l (Zwank et al., 2002, DeVaull et al., 2003), when MTBE oxygenated gasoline comes in contact with
water. As such, it is difficult to determine the source of TBA at MTBE contaminate sites. The USEPA
(2005) cautions:
“TBA accumulation by itself is not convincing evidence of MTBE biodegradation. This makes it
particularly difficult to use conventional monitoring data to document biodegradation of MTBE at
field scale, or to extract rate constants for attenuation that can be used in predictions of future
behavior of plumes.”
High TBA concentrations, relative to MTBE concentrations, at contaminated sites can be a result of the
differences in the physical/chemical properties of the two chemicals and how that influences their fate in
the subsurface, especially in the presence of remediation activity on site. TBA is more water-soluble than
MTBE, and has a comparatively higher solubility out of gasoline (compared to MTBE). At the same time,
TBA is much less volatile than MTBE (see Table 5.1 and Figure 5.3). Since most gasoline releases occur
above groundwater, significant mass of contaminants can reside in the unsaturated zone, on its way to
groundwater. Fluctuations in the water table can further “smear” the contaminants in the unsaturated zone.
Above the water table, contaminants are subject to volatilization to the atmosphere, as well as dissolution
and transport to the groundwater through precipitation (rain) events:
“If there are strong seasonal effects on the recharge of precipitation to ground water, these effects
may be reflected in the measured concentrations of MTBE in monitoring wells. As the water table
moves up and down in response to recharge, the water may wet more or less of the gasoline in the
smear zone, resulting in higher or lower concentrations of MTBE in the well. If a plume is
vertically heterogeneous, the screened interval of a monitoring well may sample different vertical
regions in the same plume as the water table shifts.” (USEPA, 2005).
For example, Beckenbach et al. (2002) noted that seasonal precipitation (infiltration) and water table
70

fluctuations significantly influence MTBE groundwater concentrations relative to BTEX concentrations.
That is expected because of the high solubility and recalcitrance of MTBE relative to BTEX. Similarly,
TBA groundwater concentrations would be expected to be influenced by infiltration and changes in
groundwater level. The often reported organic carbon partitioning coefficient Koc for TBA of 37 is similar
to that of benzene and much greater than that of MTBE (Table 5.1 and Figure 5.3; ITRC, 2005;
Kavanaugh and Stocking in USEPA Blue Ribbon Panel http://www.epa.gov/air/caaac/mtbe/mtbeethan.pdf).
However, Schmidt et al. (2004) computed a Koc for TBA of 2.6, more in line with other alcohols. For this
value of Koc, sorption of TBA would tend to be negligible in most aquifers.

Table 5.1. Properties of TBA, MTBE, benzene, toluene, and p-xylene.

Compound

Solubility
(mg/l)

Vapor Pressure
(mm mercury)

Henry’s Law
(dimensionless)

TBA

Misciblea

40-42a

4.8 x 10-4 – 5.9 x 10-4 a 0.150.33

2.6, 37

0.023 – 0.123a

15.5

11

350

36 - 141

MTBE
Benzene

43,000-54,000a
a

1,780
b

245-256a
76.0-95.2
b

a

0.22

a

Kfw

Koc

b

p-Xylene
589
8.9
0.283
4350
159
a
API, 2000
b
Eweis et al., 1998
Kfw from Zwank et al., 2002.
Koc for TBA of 37, MTBE and benzene from ITRC (2005).
Koc for TBA of 2.6 from Schmidt et al. (2004).
Koc for p-xylene based on mixed isomers: http://www.epa.gov/oppsrrd1/REDs/xylene_red.pdf.

71

Figure 5.3. Physical/chemical properties related to the fate and transport of selected oxygenated gasoline constituents including MTBE, TBA and benzene. The
physical/chemical properties of MTBE and TBA are considerably different such that they are expected to naturally separate in the environment after a MTBE
reformulated gasoline release. (Figures from Fate and Transport of Ethanol in the Environment presented M.C. Kavanaugh and A. Stocking to the USEPA Blue
Ribbon Panel, http://www.epa.gov/air/caaac/mtbe/mtbeethan.pdf)

72

Table 5.2: State ratings of SVE and bioremediation used to remediate MTBE contaminated soil
(NEIWPCC, 2003)
Very good
Good
Moderately good
Poor
12
3
10
SVE
3
6
12
Biodegradation

Remediation activity and natural processes such as volatilization into the atmosphere, can also influence,
and significantly change, the mass of contaminants in the unsaturated zone, available to impact
groundwater (Eweis et al., 2007). SVE is aimed at “vacuuming” volatile contaminants from the unsaturated
zone and has been proven very effective at locally reducing MTBE concentrations in soil. It is the
technology most commonly used for treating soil contaminated with MTBE (USEPA, 2004). In fact, in a
survey of state experiences with remediation process for MTBE and other oxygenate contamination at
LUST sites (NEIWPCC, 2003), soil vapor extraction (SVE)21 was the most widely cited technology for
treating contaminated soils. Furthermore, in terms of performance, more states rated SVE on the “very
good” range of the spectrum, while biodegradation was rated more on the “poor” end of the spectrum
(Table 5.2).
As MTBE mass is removed through SVE, less is available to impact groundwater. Conversely, TBA, being
much less volatile from water, would likely remain behind, facilitating its migration to the water table
(Eweis et al., 2007). This change in contaminant mass (mainly MTBE loss) in the unsaturated zone is
expected to be reflected in groundwater, resulting in a rising TBA to MTBE ratio that can mirror what
might be expected from biodegradation. Figure 5.4 shows the apparent effects of dual phase extraction
(DPE, which is a combination of vapor extraction and groundwater extraction) on MTBE and TBA
concentrations in the groundwater monitoring wells.

21

Soil vapor extraction (SVE) is a process wherein vapors are pumped from wells screened in the vadose

zone above the water table
73

Figure 2.3. Location of the trial sites and other known MTBE contamination sites in the State, in relation to
the location of the private wells that have reported groundwater sampling analysis under the Private Well
Testing Act (as per the most recent data from the State). Note that the New Jersey has an estimated
400,000 private well, and approximately 20% of those wells are plotted.

17