ENVIRONMENTAL ENGINEERING SCIENCE
Volume 20, Number 5, 2003
© Mary Ann Liebert, Inc.
N-Nitrosodimethylamine (NDMA) as a Drinking Water
Contaminant: A Review
William A. Mitch,1Jonathan O. Sharp,2R. Rhodes Trussell,3Richard L. Valentine,4
Lisa Alvarez-Cohen,2and David L. Sedlak2,*
1Department of Chemical Engineering
Environmental Engineering Program
New Haven, CT 06520
2Department of Civil and Environmental Engineering
University of California, Berkeley
Berkeley, CA 94720
3Montgomery Watson Harza, Inc.
Pasadena, CA 91101
4Department of Civil and Environmental Engineering
University of Iowa
Iowa City, IA 52242
N-Nitrosodimethylamine (NDMA) is a member of a family of extremely potent carcinogens, the N-ni-
trosamines. Until recently, concerns about NDMA mainly focused on the presence of NDMA in food,
consumer products, and polluted air. However, current concern focuses on NDMA as a drinking water
contaminant resulting from reactions occurring during chlorination or via direct industrial contamination.
Because of the relatively high concentrations of NDMA formed during wastewater chlorination, the in-
tentional and unintentional reuse of municipal wastewater is a particularly important area of concern. Al-
though ultraviolet (UV) treatment can effectively remove NDMA, there is considerable interest in the de-
velopment of less expensive alternative treatment technologies. These alternative technologies include
approaches for removing organic nitrogen-containing NDMA precursors prior to chlorination and the use
of sunlight photolysis, and in situ bioremediation to remove NDMA and its precursors.
Key words: N-nitrosodimethylamine (NDMA); nitrosation; unsymmetrical dimethylhydrazine (UDMH);
chlorination; photolysis; bioremediation
N-NITR OSODIM ETHY LAMINE (NDMA) IS A MEMB ER of a
family of extremely potent carcinogens, the N-ni-
trosamines (U.S. EPA, 2002). Their cancer potencies are
much higher than those of the trihalomethanes. Much of
the recent focus on NDMA as a drinking water contami-
nant can be traced to the detection of NDMA in drinking
water wells near a rocket engine testing facility in Sacra-
mento County, CA, that used unsymmetrical dimethylhy-
*Corresponding author: Department of Civil and Environmental Engineering, University of California, Berkeley, Berkeley, CA
94720. Phone: 510-643-0256; Fax: 510-642-7483; E-mail: email@example.com
drazine (UDMH)-based rocket fuel. With groundwater
NDMA concentrations as high as 400,000 ng/L on site
and 20,000 ng/L off site, it became necessary to close
downgradient drinking water wells (DHS, 2002; Mac-
Donald, 2002). The U.S. EPA established a cleanup level
of 0.7 ng/L for NDMA in groundwater (U.S. EPA, 2001),
based on a risk assessment target of an increased lifetime
cancer risk of 1026in drinking water (U.S. EPA, 2002).
The subsequent discovery of NDMA at concentrations up
to 3,000 ng/L downgradient of another rocket engine test-
ing facility in the San Gabriel Valley (CA) spurred the
California Department of Health Services to sponsor a sur-
vey of NDMA in California drinking waters (DHS, 2002).
The results of this survey demonstrated that NDMA
occurrence was not limited to regions proximal to facil-
ities that used UDMH-based fuels. Rather, NMDA de-
tected at other sites also appeared to be associated with
chlorine disinfection of water and wastewater. Especially
in locations where chlorinated wastewater effluent was
used for aquifer recharge, NDMA was present at elevated
concentrations. For example, two drinking water pro-
duction wells, under the influence of recharge water from
the advanced wastewater treatment system of the Orange
County Water District’s Water Factory 21, suspended op-
erations due to the presence of NDMA in 2000 (OCWD,
2000a). Groundwater injection of treated wastewater
from Water Factory 21 was reduced from 7 to 1 million
gallons per day pending the installation of an expensive
ultraviolet treatment system to remove the NDMA
prior to injection (OCWD, 2000b). Even more recently,
NDMA was detected in treated drinking water from
sources that were not impacted by wastewater effluent or
industrial sources, especially when monochloramine was
used to maintain a chlorine residual (DHS, 2002).
Although NDMA is listed as a priority pollutant (CFR,
2001), a federal maximum contaminant level (MCL) has
not been established for drinking water. Moreover,
NDMA is not even on the Candidate Contaminant List,
which sets the priorities for future regulation of drinking
water (U.S. EPA, 1998). However, other regulatory agen-
cies have established NDMA guidelines. The Ontario
Ministry of the Environment and Energy established an
Interim Maximum Acceptable Concentration of 9 ng/L
for NDMA (MOE, 2000). After discovering the wide-
spread presence of NDMA, the California Department of
Health Services set an interim action level of 20 ng/L,
which was later reduced to 10 ng/L (DHS, 2002).
Despite heightened recent concern, NDMA is not re-
ally an emerging contaminant. Since the 1960s, toxicol-
ogists have studied the health effects of nitrosamines. The
concern focused on their widespread occurrence in food
and consumer products, particularly beer, meats cured
with nitrite, tobacco smoke, and rubber products includ-
ing baby bottle nipples (IARC, 1978). Concerns about
human exposure to NDMA from industrial sources also
were voiced previously. During the 1970s, NDMA was
detected in the air and water adjacent to a factory near
Baltimore (MD) that produced UDMH from NDMA
(Shapley, 1976; Fine et al., 1977; Fine, 1978). More
alarming was the detection of NDMA in the air upwind
of the plant in downtown Baltimore (0.1 mg/m3), and at
other sites in Belle, WV (0.1 mg/m3), and New York City
(0.8 mg/m3), areas with no known industrial sources of
NDMA. On the basis of those observations, some re-
searchers suggested that NDMA formed in the polluted
atmosphere could be responsible for elevated urban can-
cer rates (Shapley, 1976). However, it was subsequently
determined that a factory using dimethylamine was lo-
cated near Belle, WV, and that air concentrations of
NDMA in areas not impacted by industrial processes were
orders of magnitude lower than the initial reports (Hanst
et al., 1977; Cohen and Bachman, 1978; Fine, 1978).
Prior review articles (IARC, 1978; ASTDR, 1989)
have focused on the occurrence and toxicology of NDMA
in food and consumer products. In addition to research
occurring after 1989, this review covers material from
prior review articles that is relevant to water treatment.
Prior to the recent interest in low-level NDMA occur-
rence, analysis of the compound usually was performed
by liquid–liquid extraction and gas chromatography/mass
spectrometry (GC/MS) or gas chromatography with a
thermionic detector. The detection limit of the method
was approximately 1,000 ng/L. The most common tech-
nique currently used for analysis of low concentrations
of NDMA involves extraction, preconcentration, and
analysis by gas chromatography with tandem mass
spectrometry in the chemical ionization mode (GC/
CI/MS/MS) or gas chromatography with high resolution
mass spectrometry (GC/HRMS). These methods typi-
cally have detection limits around 1 ng/L. Although a
standard method for low-level quantification of NDMA
has not been published, several methods have been shown
to yield accurate and reproducible results. In the first step,
residual chlorine in the sample is quenched with ascor-
bic acid or sodium thiosulfate to prevent an artifact due
to reaction of chlorine with methylene chloride to form
NDMA (Cohen and Bachman, 1978). Deuterated NDMA
is added for use in isotope dilution to reduce the uncer-
tainty associated with extraction efficiency. In some
methods, the sample is extracted in methylene chloride
by the separatory funnel method according to U.S. EPA
Method 3510C (U.S. EPA, 1996). Unfortunately, this
390 MITCH ET AL.
method yields low recoveries, and may generate difficult-
to-handle emulsions when used for wastewater effluent
samples. Extraction efficiencies can be improved to ap-
proximately 50% by the addition of up to 100 g/L of so-
dium chloride (Yoo et al., 2000). Other methods employ
continuous liquid–liquid extraction via U.S. EPA Method
3520C (U.S. EPA, 1998), which involves extraction of
the sample with 100–300 mL methylene chloride for ap-
proximately 6–18 h. Continuous liquid–liquid extraction
avoids problems associated with emulsions in wastewater
samples, and can yield extraction efficiencies of up to
60% (Mitch et al., 2003). The methylene chloride ex-
tracts are then concentrated to 1 mL or less using rotary
evaporators or nitrogen blowdown.
There have been several attempts to use solid-phase
extraction to improve extraction efficiency and to reduce
the volume of methylene chloride required for extrac-
tions. Jenkins et al. (1995) reported a NDMA solid-phase
extraction method involving the use of carbonaceous
Ambersorb 572 resin that reduces the volume of meth-
ylene chloride required to 400 mL. Recoveries were ap-
proximately 30%. This method suffers from difficulties
arising from fragmentation of the resin and subsequent
recovery for extraction (Tomkins and Griest, 1996).
Tomkins and Griest (1996) described solid phase extrac-
tion using a carbon-based Empore disk that resulted in
60% recovery. Unfortunately, the carbon-based Empore
extraction disks are no longer available.
Following extraction, NDMA is separated by capillary
gas chromatography (most often with DB-5 or DB-1701
columns) followed by detection by one of several meth-
ods: thermal energy analyzers (Fine et al., 1977; Kimoto
et al., 1981), chemiluminescent nitrogen detectors (Tomkins
et al., 1995), and high-resolution mass spectrometers
(Taguchi et al., 1994). Although these methods are still
used, the most common method used for determination
of low concentrations of NDMA involves the use of
chemical ionization followed by tandem mass spec-
Analytical methods also have been described for
NDMA precursors. Mitch et al. (2003) describe methods
for analyzing total concentrations of the organic nitrogen
precursors for NDMA formation during chlorination of
water and wastewater. Although several methods for the
detection of dimethylamine (and other primary and sec-
ondary amines) have been previously described (Scully
et al., 1988; Hwang et al., 1994, 1995; Lopez et al., 1996;
Sacher et al., 1997; Abalos et al., 1999; Liu et al., 2001),
a method recently was developed specifically for use in
water or wastewater (Mitch et al., 2003).
Obtaining blanks that are free from NDMA and
NDMA precursors can be problematic. Deionized or dis-
tilled water can be contaminated with several ng/L of
NDMA (Kimoto et al., 1981) as well as NDMA precur-
sors (Gerecke and Sedlak, 2003). A UV lamp can be used
to destroy NDMA in deionized water prior to use or
NDMA-free water can be obtained by purchase of HPLC
SOURCES AND OCCURRENCE
NDMA can be released directly from industrial sources
as a contaminant of products such as liquid rocket fuel,
or it can be formed in solution from chemical reactions.
Available data suggest that there are two major pathways
for NDMA formation: (1) nitrosation, and (2) formation
by oxidation of UDMH. Although the two pathways dif-
fer in their mechanisms of formation, the organic nitro-
gen precursors involved in both reactions may be identi-
Nitrosation: NDMA formation via nitrite
Nitrosation involves the formation of nitrosyl cation or
similar nitrogen-containing species, such as dinitrogen
trioxide (N2O3), during acidification of nitrite (reactions
1 and 2; Mirvish, 1975). The nitrosyl cation then reacts
with an amine, such as dimethylamine, to form NDMA.
This reaction occurs most rapidly at pH 3.4, reflecting a
balance between the protonation of nitrite (pKaof
HNO253.35) and the increased fraction of dimethy-
lamine in the reactive, deprotonated from with increas-
ing pH (pKaof H2N(CH3)2
NO11(CH3)2NH (CH3)2N—N 5O1H1(2)
This nitrosation mechanism is believed to be respon-
sible for the observed formation of NDMA in vegetables,
fish, and especially meat products cured with nitrite to
prevent the growth of Clostridium botulinum, the bac-
terium that generates botulism toxin (IARC, 1978). Ni-
trate also can contribute to nitrosation because it can be
reduced to nitrite by bacteria in the mouth (Preussmann,
1984). In vivo nitrosation occurs when nitrite enters the
acidic environment of the stomach (Shapley, 1976). Nu-
cleophilic anions, particularly thiocyanate (a constituent
of saliva), enhance the rate of nitrosation through cat-
alytic NDMA formation from nitrite (Fan and Tannen-
baum, 1973). Although the U.S. Food and Drug Admin-
istration reduced the concentrations of nitrite allowed for
curing meat to a maximum of 120 ppm, several meat
processors add reducing agents such as ascorbic acid to
quench nitrosating agents and minimize in vivo NDMA
formation (Preussmann, 1984).
Although the rate of nitrosation is slow at neutral and
NDMA AS A DRINKING WATER CONTAMINANT 391
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
basic pH, several studies indicate that photochemical re-
actions (Ohta et al., 1982), formaldehyde (Keefer and
Roller, 1973) and fulvic acid (Weerasooriya and Dis-
sanayake, 1989) can catalyze nitrosation at circumneu-
tral pH. For example, Ayanaba and Alexander (1976) ob-
served that the addition of high concentrations of nitrite
to lake water containing dimethylamine resulted in the
formation of NDMA. Although the concentration of
NDMA increased as pH decreased, NDMA formation
was observed even at pH 6.2. Denitrifying bacteria that
colonize the organs of sick patients also are known to
catalyze the nitrosation reaction at circumneutral pH val-
ues (Leach et al., 1987). Furthermore, catalysis of nitro-
sation by the yeast Candida albicans in the mouth is
thought to be potentially responsible for some oral can-
cers (Krogh et al., 1987).
Reactions similar to nitrosation also can occur in the
atmosphere. Combustion often results in the formation of
nitrogen-containing species (i.e., NOxor nitroso radicals)
that react with species such as dimethylamine to produce
nitrosamines. Relatively slow NDMA formation may oc-
cur in the atmosphere due to reaction of NOx(principally
atmospheric nitrous acid) with dimethylamine released
by industrial sources (Hanst et al., 1977; Cohen and
Bachman, 1978). However, atmospheric NDMA may ac-
cumulate only at night because it degrades quickly by
sunlight photolysis (Hanst et al., 1977; Tuazon et al.,
Gas-phase nitrosation may explain the occurrence of
NDMA in cigarette smoke, malt beverages, dried foods,
and rubber products (Preussmann, 1984; ATSDR, 1989).
NDMA can be formed during food drying or during the
barley malting process as a result of NOxin the exhaust
of air heaters (Preussmann, 1984; ATSDR, 1989; Sen et
al., 1996). For example, concentrations of NDMA de-
tected in beer dropped by nearly an order of magnitude
when malting houses switched to the use of indirect
heaters to prevent contact of the barley with heater ex-
haust or when they applied sulfur dioxide to the flue gas
as a quenching agent. Prior to these modifications, beer
was estimated to be the major dietary contributor to daily
NDMA ingestion. Occupational exposures to NDMA are
high in the tire and rubber industries where nitroso rad-
icals in engine exhaust react with amine-containing ac-
celerators used for vulcanization (Preussmann, 1984).
NDMA also was observed in the wastewater effluents of
a variety of industrial plants manufacturing amines, her-
bicides, pesticides, pharmaceuticals (Cohen and Bach-
man, 1978) and rubber. For example, NDMA has been
detected at concentrations up to 2 mg/L in the wastewater
effluent of a tire factory in Ontario (Ash, 1995).
Nitrosation reactions in food and consumer products
may represent a significant exposure source. The esti-
mated daily intake of NDMA for an average German diet
was 0.2 mg/day (Tricker et al., 1994). The most impor-
tant dietary source of NDMA may be preserved meat and
fish products, beer, and tobacco (Fine, 1978; ATSDR,
1989). NDMA is occasionally detected in cheese and bak-
ery products (Uibu et al., 1978) as the result of the dry-
ing process for cheeses or catalysis of nitrosation reac-
tions by yeast. The formation of NDMA in amine-
containing toiletry and cosmetic products, such as sham-
poo, has been attributed to the use of nitrite (Spiegel-
halder and Preussmann, 1984). NDMA volatilizing from
upholstery also is detectable in the interior air of auto-
mobiles (ATSDR, 1989).
Unsymmetrical dimethylhydrazine (UDMH)
oxidation: NDMA formation from chlorine and
During the 1980s, the formation of NDMA was re-
ported when hypochlorite was used to treat waste
UDMH-containing rocket fuel (Brubaker et al., 1985,
1987). NDMA also has been observed as a byproduct of
UDMH oxidation by cupric ion (Banerjee et al., 1984),
potassium permanganate, iodate (Castegnaro et al.,
1986), hydrogen peroxide, and oxygen (Lunn et al., 1991;
Lunn and Sansone, 1994). The formation of NDMA from
oxidation of UDMH is maximized at neutral and high pH
(Lunn et al., 1991).
The formation of NDMA during water chlorination
was reported in laboratory experiments in 1980 (Kimoto
et al., 1980, 1981). The formation of NDMA was later
documented after chlorination at full-scale drinking wa-
ter treatment plants and at wastewater treatment plants
(Jobb et al., 1994; Ash, 1995; Child et al., 1996). Be-
cause NDMA is formed when UDMH is oxidized, any
chlorination reactions that produce UDMH also should
produce NDMA. The reaction between monochloramine
and dimethylamine to form UDMH (Yagil and Anbar,
1962) and the reaction of monochloramine with trimethy-
lamine to form a 1,1,1-trimethyl hydrazinium salt (Omi-
etanski and Sisler, 1956) have been known for some time.
Delalu et al. (1981; Delalu and Marchand, 1987, 1989a,
1989b) described the kinetics of the formation of UDMH
from the reaction of monochloramine and dimethylamine
and the subsequent oxidation of UDMH at high concen-
trations of reactants. However, they did not attempt to
measure the formation of NDMA.
Until recently, NDMA formation during chlorination
was assumed to occur via the nitrosation pathway (Ki-
moto et al., 1981; Jobb et al., 1994; OCWD, 2000b).
However, Mitch and Sedlak (2002a) and Choi and Valen-
tine (2002a, 2002b) demonstrated that NDMA formation
during chlorination could occur through UDM H as an in-
392 MITCH ET AL.
termediate (Fig. 1). The rate of UDMH formation via this
process increases with pH (Yagil and Anbar, 1962). Al-
though stable at high pH, the UDMH intermediate is ox-
idized nearly instantaneously at circumneutral pH to form
NDMA at low yields (,1%) (Mitch and Sedlak, 2002a).
The maximum rate of formation occurs at circumneutral
pH. Due to the preliminary slow step, the overall rate of
formation is extremely slow, resulting in formation of
NDMA over a period of days. Mitch and Sedlak (2002b)
demonstrated that chloramination of amines via this path-
way was a plausible explanation to account for the for-
mation of NDMA during wastewater chlorination.
However, it was unclear whether dimethylamine con-
centrations in secondary effluent are sufficient to account
for NDMA formation or whether other organic nitrogen
compounds are more important precursors.
Bromide ion is frequently a trace component of drink-
ing water and wastewater. It is readily oxidized by free
chlorine and monochloramine although the rate of reac-
tion with monochloramine is several orders of magnitude
slower (Trofe et al., 1980). In the presence of excess am-
monia, bromamines are readily produced when hypochlo-
rite is added to water. Given the similarity of bromamine
to chloramine chemistry and the generally increased re-
activity of bromamines compared to chloramines, it is not
surprising that Choi and Valentine (2002b, 2002c) ob-
served a catalytic effect of bromide on NDMA forma-
The two-step mechanism elucidated by Mitch and Sed-
lak (2002a) and Choi and Valentine (2002a, 2002b) is
consistent with observations (Najm and Trussell, 2001;
Berger et al., 2002; Najm and Ma, 2002; Wilczek et al.,
2002) that the use of monochloramine in water treatment
greatly increases NDMA formation. Because NDMA for-
mation is slow (Mitch and Sedlak, 2002a; Mitch et al.,
2003), the use of monochloramine to maintain a chlorine
residual can result in increasing concentrations of NDMA
within the distribution system. However, the problem is
not restricted to chloramines. In the absence of ammo-
nia, hypochlorite also can produce NDMA through reac-
tion with secondary amines, but the rate of formation is
approximately an order of magnitude lower than that ob-
served with monochloramine (Mitch and Sedlak, 2002a).
Several studies have documented NDMA formation
in waters treated with ion-exchange resins having qua-
ternary amine functional groups that could serve as
NDMA precursors (Fiddler et al., 1977; Gough et al.,
1977; Kimoto et al., 1980). In many of these studies, a
chlorine or monochloramine residual in the water prob-
ably reacted with resin functional groups. However,
Najm and Trussell (2001) found that even distilled wa-
ter leached significant concentrations of NDMA (up to
approximately 60 ng/L after 4 h of contact) from anion
exchange resins. These concentrations doubled in the
presence of 1 mg/L nitrite, suggesting a mechanism
other than UDMH oxidation was involved. Although
NDMA may form from the reaction of hypochlorite
with amine-containing polymers used in water treat-
ment, two studies indicated that the concentrations
formed are not likely to be significant under normal
drinking water conditions (Child et al., 1996; Najm and
NDMA AS A DRINKING WATER CONTAMINANT 393
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
Figure 1. Pathway for NDMA formation during chloramination of dimethylamine via a UDMH intermediate (Mitch and Sed-
NDMA OCCURRENCE IN DRINKING
WATERS, WASTEWATERS, AND
The discovery of elevated concentrations of NDMA in
treated drinking water in Ohsweken, Ontario, in 1989
prompted a survey of 145 Ontario drinking water plants
(Jobb et al., 1994; MOE, 1998). This survey indicated that
the NDMA concentration in the treated water from most
plants was less than 5 ng/L, although some samples ex-
ceeded 9 ng/L. Similar results were obtained by the Cal-
ifornia Department of Health Services during a survey of
NDMA concentrations in drinking water systems con-
ducted in 2001 (DHS, 2002). The results showed that 3
of the 20 chloraminated supplies surveyed contained
NDMA concentrations greater than 10 ng/L, while none
of the eight water supplies that used only free chlorine
disinfection exhibited NDMA levels above 5 ng/L. One
of the four drinking water supplies surveyed that employ
anion exchange treatment also showed NDMA concen-
trations in excess of 10 ng/L. Other sampling programs
confirmed that the majority of treated drinking water sam-
ples contain less than 10 ng/L of NDMA (Tomkins et al.,
1995; Tomkins and Griest, 1996). No sampling was con-
ducted within water distribution systems, although the
concentrations are anticipated to gradually increase in the
presence of chlorine or chloramine residuals.
In contrast to results from drinking water treatment
plants, effluents from conventional and advanced waste-
water treatment plants contain relatively high concentra-
tions of NDMA. NDMA is often present in raw sewage
prior to chlorination. For example, NDMA concentra-
tions as high as 105,000 ng/L have been reported in ef-
fluents from printed circuit board manufacturers using
NDMA-contaminated dimethyldithiocarbamate to re-
move metals (OCSD, 2002). These industrial inputs re-
sulted in concentrations of NDM A of approximately
1,500 ng/L in raw sewage. As a result of removal pro-
cesses that occur during secondary treatment, NDMA
concentrations in unchlorinated secondary effluent often
are less than 20 ng/L, although industrial inputs can lead
to large spikes in NDMA influent and effluent concen-
Chlorination of secondary wastewater effluent typi-
cally results in the formation of between 20 and 100 ng/L
NDMA (Mitch and Sedlak, 2002b). Consistent with the
UDMH-intermediate mechanism, nitrification of waste-
water to completely remove ammonia prior to hypochlo-
rite addition reduces NDMA formation by approximately
an order of magnitude (Mitch and Sedlak, 2002e). In
wastewater recycling plants receiving secondary waste-
water effluent, NDMA concentrations in microfiltration
effluent may increase by approximately 30–50 ng/L as a
result of chlorination before the membrane to prevent bi-
ological growth (L. McGovern, personal communica-
tion). NDMA also has been detected in dried municipal
sewage sludge used for agricultural fertilizer, but in this
case the formation pathway may be biologically medi-
ated nitrosation during anaerobic digestion (Brewer et al.,
1980; ATSDR, 1989).
ORGANIC NITROGEN PRECURSORS:
SOURCES AND OCCURRENCE
Both of the NDMA formation mechanisms involve
reactions between an inorganic, nitrogen-containing
species (e.g., N2O3, NH2Cl) and an organic nitrogen
species. Not surprisingly, dimethylamine has been dem-
onstrated to be the most effective organic nitrogen pre-
cursor of NDMA formation by both the nitrosation path-
way (Fiddler et al., 1972) and the UDMH pathway (Mitch
and Sedlak, 2002b). Both pathways also can produce
NDMA from tertiary amines containing dimethylamine
functional groups (e.g., trimethylamine and dimethy-
lethanolamine) but at lower yields. Nitrosation of
trimethylamine-N-oxide, a common constituent of urine
(Zuppi et al., 1997), resulted in some NDMA formation,
but much less than did trimethylamine (Fiddler et al.,
1972). Many other organic nitrogen-containing mole-
cules, including the primary amine monomethylamine,
the quaternary amine tetramethylamine, and amino acids
or proteins, did not form significant concentrations of
NDMA after chloramination (Mitch and Sedlak, 2002b).
Fiddler et al. (1972) found that nitrosation of quaternary
amines that contained trimethylamine functional groups
resulted in four orders of magnitude lower NDMA con-
centrations than did trimethylamine. Significant organic
nitrogen precursors for NDMA formation therefore ap-
pear to be limited to dimethylamine and tertiary amines
with dimethylamine functional groups.
The lower yields of NDMA from species other than
dimethylamine are not unexpected given the need to
break a C—N bond prior to NDMA formation. In the
case of the nitrosation pathway, Ohshima and Kawabata
(1978) described a complex reaction scheme for NDMA
formation from trimethylamine-N-oxide and trimethy-
lamine that accounts for the dealkylation required to form
the dimethylamine portion of NDMA using the proposed
pathways of earlier researchers such as Smith and
Loeppky (1967) and Lijinsky and Singer (1975). A deal-
kylation scheme for NDMA formation during chlorami-
nation of tertiary amines containing dimethylamine func-
tional groups may involve chlorine transfer to the
nitrogen atom followed by elimination of HCl to form an
iminium ion (Ellis and Soper, 1954). Hydrolysis of the
394 MITCH ET AL.
iminium ion results in formation of the secondary amine.
Finally, Mitch and Sedlak (2002b) demonstrated that
chloramination of other secondary amines or tertiary
amines containing functional groups other than dimethy-
lamine resulted in the formation of their respective ni-
trosamines in quantities similar to those associated with
NDMA formation from dimethylamine and trimethy-
Dimethylamine is present in food, and can be liberated
from food during digestion (Tricker et al., 1994). Nitro-
gen-containing organic molecules, such as the cell mem-
brane structural lipid phosphatidyl choline (lecithin) and
amino acids, are broken down by bacterial flora in the
gastrointestinal tract to trimethylamine (Simenhoff et
al., 1976; Tricker et al., 1994). After absorption into the
bloodstream, a portion of the trimethylamine is demethy-
lated to dimethylamine and excreted via the urine, gas-
tric juice, or bile. The remainder is oxidized to trimethy-
lamine-N-oxide, which is excreted in the urine in
concentrations usually twice as high as those of dimethy-
lamine (Zuppi et al., 1997). Tricker et al. (1994) found
that dimethylamine is present in human urine (average
concentration is approximately 40 mg/L) and feces (av-
erage concentration is 0.41 mg/mL). Dimethylamine
also has been detected in the feces of dairy cattle (van
Amines also are produced outside of the body by mi-
crobes via the vitamin B6-mediated degradation of amino
acids (Metzler, 1977). Ayanaba and Alexander (1974)
demonstrated that addition of relatively high concentra-
tions of trimethylamine or tetramethylthiuram disulfide
(thiram) to lake water or municipal sewage resulted in
the microbiological production and eventual consump-
tion of dimethylamine. Trimethylamine-N-oxide is pres-
ent in seafood, and may be broken down to trimethy-
lamine by bacteria (Ohshima and Kawabata, 1978).
As a result of excretion and industrial activities, di-
methylamine concentrations in primary wastewater ef-
fluent typically range from 20 to 80 mg/L (Mitch and
Sedlak, 2002c). Dimethylamine is readily degraded by
bacteria. As a result, concentrations in secondary waste-
water effluents are generally low (i.e., average 54 mg/L;
Mitch and Sedlak, 2002c). Mitch and Sedlak (2002c) con-
cluded that dimethylamine could only account for ap-
proximately 10% of the NDMA formed when secondary
wastewater effluent was chloraminated. However, other
authors, using a less sensitive HPLC method, have found
no significant loss of dimethylamine upon secondary bi-
ological treatment (Hwang et al., 1995; Abalos et al.,
In unpolluted waters, dimethylamine concentrations
are generally less than 0.1 mg/L (Gerecke and Sedlak,
2003). These concentrations can not account for NDMA
formation during chlorination (Mitch et al., 2003). How-
ever, dimethylamine, methylamine and morpholine were
detected at concentrations up to 3 mg/L in the Rhine and
Elbe Rivers in Germany (Sacher et al., 1997) where the
input of wastewater effluents may be significant. Under
these conditions, dimethylamine from unintentional reuse
of municipal wastewater effluent could be an important
Resins used in water and wastewater treatment also
may be sources of dimethylamine and other organic ni-
trogen-containing NDMA precursors. NDMA itself may
be a contaminant of carbonaceous resins and activated
carbon at levels up to approximately 10 mg/kg (Kimoto
et al., 1981). Najm and Trussell (2001) found that ex-
traction of strong-base anion-exchange resins containing
dimethyl-ethanol or trimethyl quaternary functional groups
with distilled water in the absence of chlorine resulted in
concentrations of NDMA up to approximately 50 ng/L.
Moreover, while NDMA was not detected in effluent
from resins containing triethyl or tripropyl functional
groups, N-nitrosodiethylamine and N-nitrosodi-n-propy-
lamine were detected, respectively (Najm and Trussell,
2001). NDMA precursors can leach from functional
groups on quaternary amine-containing exchange resins
(Cohen and Bachman, 1978; Kimoto et al., 1980; Najm
and Trussell, 2001). At elevated temperatures (i.e., 78°C),
some of the quaternary amine functional groups on resins
demethylate to form trimethylamine (Fiddler et al.,
Resins and granular activated carbon may promote
NDMA formation by surface-catalyzed reactions. Ange-
les et al. (1978) suggested that mixed bed resins promote
nitrosation of precursors because proton displacement on
cationic resins creates acidity, which promotes nitrosa-
tion reactions on adjacent anionic resins (usually con-
taining amine precursors) to which nitrite may adsorb.
An enhancement of nitrosation reactions was observed
when a nitrifying biofilm was active on granular activated
carbon (DiGiano et al., 1986), possibly as a result of bi-
ological catalysis. However, NDMA formation by this
pathway is unlikely to be significant under typical drink-
ing water treatment conditions.
Other industrial products containing dimethylamine
functional groups that could serve as precursors include
fungicides such as thiram (tetramethylthiuram disulfide)
(IARC, 1978) Graham et al., 1995), pesticides, and her-
bicides such as 2,4-D, which are formulated as a di-
methylamine salts (Fine, 1978; Child et al., 1996), drugs
such as ranitidine (IARC, 1978), and amine-containing
accelerators for vulcanization of tires (Fig. 2). The dithio-
carbamates, which are a family of compounds used as
fungicides, herbicides, and as chelating agents to remove
cationic metals from industrial wastewater, contain
NDMA AS A DRINKING WATER CONTAMINANT 395
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
readily hydrolyzable dimethylamine functional groups
(Weissmahr and Sedlak, 2000).
Amine-based polymers (Child et al., 1996; MOE,
1998; Najm and Trussell, 2001) and unknown contami-
nants of alum (Jobb et al., 1994) also have been identi-
fied as sources of NDMA precursors. In the case of
wastewater treatment, treatment polymers containing di-
methylamine groups become associated with particles,
and can form a significant fraction of the total organic
nitrogen precursors during wastewater treatment (S. Carr,
Removal of NDMA
The vapor pressure of NDMA is estimated to be rela-
tively high at 360 Pa at 20°C (Klein, 1982). Due to the high
water solubility of NDMA, the estimated Henry’s Law con-
stant for NDMA is low at 2.6 31024atm M2120°C
(ATSDR, 1989; Mirvish et al., 1976). Therefore, volatiliza-
tion from natural waters and air stripping are unlikely to
result in significant removal of NDMA from solution. As
a small, uncharged molecule, NDMA is poorly removed
via reverse osmosis membranes. Within wastewater recy-
cling plants, NDMA was removed with approximately 50%
efficiency by thin-film composite reverse osmosis mem-
branes (L. McGovern, personal communication).
Due to the presence of polar functional groups, NDMA
is hydrophilic, with a log Kow value of 20.57 (ATSDR,
1989). As a result, NDMA sorbs poorly to soil, activated
carbon, and other sorbents. Bituminous coal granular-ac-
tivated carbon was used in interceptor trenches to remove
NDMA arising from groundwater contamination at the
Rocky Mountain Arsenal (Fleming et al., 1996). Flem-
ing et al. (1996) found that sorption onto hydrophilic sor-
bents such as silica, acrylic resins, and zeolite were in-
significant. Ambersorb 572 carbonaceous resin was
found to be the most effective sorbent, followed by co-
conut shell carbon. However, the Freundlich isotherm K
and 1/nconstants for Ambersorb 572 (9.65 31023mg/g
and 1.17, respectively) were low, which resulted in pro-
hibitive treatment costs. Consistent with these observa-
tions, the transport of NDMA was not retarded through
soil columns (Dean-Raymond and Alexander, 1976).
Currently, the most commonly applied aqueous NDMA
treatment method is photolysis by ultraviolet (UV) radi-
ation. NDMA absorbs light strongly between 225 and 250
nm (Fig. 3; lm ax 5228 nm where «57380 M21cm21).
This wavelength is at the lower end of the transparency
of water to UV. The absorption results in a pto p*tran-
sition (Polo and Chow, 1976; Stefan and Bolton, 2002)
followed by cleavage of the N—N bond, most likely via
hydrolysis to dimethylamine and nitrous acid, or by form-
ing nitroso and dimethylamine radicals (reaction 3).
The major products of the reaction are dimethylamine
and nitrite, while minor products include nitrate, for-
396 MITCH ET AL.
Figure 2. Industrial products that could be precursors for NDMA formation: the pesticide tetramethylthiuram disulfide (thiu-
ram), the fungicide, herbicide, and metal chelator dimethyldithiocarbamate (DTC), and the H-2 receptor antihistamine pharma-
ceutical ranitidine (Zantac).
O N N O N 1 N
maldehyde, formate, and perhaps methylamine (Stefan
and Bolton, 2002). The quantum yield at pH 7 is 0.13
(Stefan and Bolton, 2002). Dimethylamine is resistant to
further photolytic reactions, while nitrite is readily oxi-
dized to nitrate. NDMA also has a secondary absorption
peak between 300 and 350 nm (lmax 5332 nm, where
«5109 M21cm21) within which excitation occurs via
an n-to-p*transition (Stefan and Bolton, 2002). This sec-
ondary peak overlaps well with the intensity of UV pro-
duced by medium-pressure mercury lamps (Stefan and
Bolton, 2002). However, whether low- or medium-pres-
sure lamp systems are more efficient for NDMA de-
struction is still unclear. Under conditions typically en-
countered in drinking water treatment systems, the UV
dosage required for a one order of magnitude decrease in
NDMA concentration is approximately 1,000 mJ/cm2,
which is approximately 10 times higher than that required
for equivalent virus removal. Therefore, UV treatment
for NDMA will be feasible but more expensive than UV
treatment for disinfection.
UV treatment has been used to remove NDMA at a
drinking water plant in Ohsweken, Ontario (Jobb et al.,
1994), in effluent from a tire factory upgradient of the
Ohsweken plant (Ash, 1995), and at Water Factory 21 in
Orange County, CA (OCWD, 2000a). Three technolo-
gies have been used for UV treatment of NDMA: low-
pressure UV lamps emitting mainly monochromatic light
at 254 nm, medium-pressure lamps emitting polychro-
matic light, and pulsed UV systems. Pulsed UV systems
have the advantage of an emission spectrum that more
closely matches the adsorption spectrum of NDMA
(Liang, 2002). However, the technology is less proven
than low- and medium-pressure UV lamps.
Because UV photolysis may not destroy NDMA pre-
cursors, some have suggested that reformation of NDMA
may occur within drinking water distribution systems if
chlorination is performed after the UV treatment (Jobb
et al., 1994). However, the concentration of dimethy-
lamine liberated when low concentrations of NDMA are
photolyzed usually will be small. If significant concen-
trations of NDMA are formed upon chlorination follow-
ing UV treatment, the formation likely results from other
NDMA precursors. Addition of hydrogen peroxide to
generate hydroxyl radical for NDMA oxidation does not
significantly increase NDMA destruction efficiency
(Jobb et al., 1994; Liang, 2002).
Photolysis of NDMA also occurs in sunlight as a re-
sult of NDMA’s secondary absorption band between 300
and 350 nm. Sunlight photolysis was used by the OCWD
as part of their initial attempt to reduce concentrations of
the compound at Water Factory 21; placing treated wa-
ter in shallow sunlit basins with residence times of ap-
proximately 1 day resulted in removal of approximately
half of the NDMA (M. Wehner, personal communi-
cation). Sunlight photolysis also may be an important
loss mechanism for NDMA applied in irrigation water
Atmospheric photolysis of NDMA removes NDMA
from the sunlit atmosphere within a few hours (Shapley,
1976; Hanst et al., 1977; Cohen and Bachman, 1978; Tu-
azon et al., 1984). NDMA is believed to decay via equa-
tion (1) with a quantum yield of 1 for l$290 nm
(Tuazon et al., 1984). The major product is dimethyl-
nitramine via the reaction of NO2with the dimethylamino
radical. Lesser products include formaldehyde and
methylnitramine. The half-life for the atmospheric reac-
NDMA AS A DRINKING WATER CONTAMINANT 397
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
Figure 3. Emission spectra for low- and medium-pressure mercury lamps, sunlight spectrum at the surface of the Earth and ab-
sorption spectrum of 1 mM NDMA.
tion of NDMA with hydroxyl radical was estimated to
be approximately 3 days compared with only 5 min for
direct photolysis (Tuazon et al., 1984).
Ozone does not appear to react with NDMA by a di-
rect mechanism either in the atmosphere (Tuazon et al.,
1984) or in water (Liang, 2002). Hydroxyl radicals pro-
duced from ozone and hydrogen peroxide can be used to
treat NDMA (Liang, 2002). However, the efficiency
of ozonation and other advanced oxidation processes
(AOPs) will be limited by the presence of hydroxyl rad-
Zero-valent iron catalyzes NDMA transformation by
hydrogenation (Gui et al., 2000; Odziemkowski et al.,
2000). This reduction reaction leads to the formation of
dimethylamine and ammonia as final products. Although
a field feasibility study conducted with canisters demon-
strated NDMA reduction, the relatively slow kinetics of
the reaction suggested that it would not be a cost effec-
tive treatment option (Cox, 2002). The addition of 0.25%
nickel to the iron increased the reduction rate by nearly
a factor of 340. However, the reaction rate for the
nickel–iron mixture decreased within 100 pore volumes.
The potential for phytoremediation of NDMA is cur-
rently unknown. However, the high aqueous solubility
of the compound is well-suited for the treatment. Let-
tuce and spinach readily took up 14 C-labeled NDMA
from irrigation water (Dean-Raymond and Alexander,
1976). 14C activity in the plants decreased with time,
suggesting that the NDMA was converted to 14CO2in
Bioremediation could hold significant potential for the
in situ treatment of NDMA contaminated water. Bacte-
rial monooxygenase enzymes may be similar to the cy-
tochrome P-450 enzymes that catalyze the NADPH-de-
pendent oxidation of NDMA in both plants and animals
(Tu and Yang, 1985; Yamazaki et al., 1992; Stiborova et
al., 2000). Mineralization (conversion to CO2) of NDMA
by undefined consortia has been observed in two studies
(Kaplan and Kaplan, 1985; Gunnison et al., 2000). Bio-
degradation also has been reported in anaerobic and aer-
obic incubations of native microbial soil consortia, with
half-lives ranging from 12 to 55 days (Tate and Alexan-
der, 1975; Oliver et al., 1979; Gunnison et al., 2000).
Biodegradation proceeded slightly faster under aerobic
conditions than under anaerobic conditions (Mallik and
Tesfai, 1981). In two cases, the NDMA biodegradation
rate may have slowed after the first few weeks of NDMA
application (Tate and Alexander, 1975; Mallik and Tes-
fai, 1981); however, these studies suffered from poor
quantification of the effect of confounding factors such
as volatilization (ATSDR, 1989). Although these studies
documented degradation intermediates, including methy-
lamine and formaldehyde, none of these studies was able
to identify the responsible micro-organisms nor elucidate
Despite the existence of NDMA-degrading bacteria
in soil, there is limited evidence for the biodegradation
of NDMA under field conditions. For example, at the
Rocky Mountain Arsenal, no significant loss of NDMA
was observed during passage through the aquifer (Gun-
nison et al., 2000). However, after groundwater was
passed through an ex situ granular activated carbon
(GAC) treatment system and reinjected into the sub-
surface, NDMA removal was observed. Because
NDMA adsorption to GAC is negligible, the GAC at
the Rocky Mountain Arsenal site may have removed
competitive substrates from solution, allowing NDMA
biodegradation to proceed. In another study, the addi-
tion of glucose or nutrient broth to microcosms hindered
NDMA mineralization, indicating that substrate com-
petition may occur (Kaplan and Kaplan, 1985). It is
likely that a complex interaction exists between dis-
solved organic nutrients necessary for the growth of
bacteria capable of degrading NDMA and a tendency
of these micro-organisms to consume these nutrients in
preference to NDMA. In addition to the lack of clear
evidence for bioremediation of NDMA in groundwater,
there is no information regarding the potential for bio-
logical removal of NDMA within drinking water treat-
ment systems such as biofiltration units.
Removal of NDMA precursors
Unlike NDMA, many nitrogen-containing NDMA
precursors, including dimethylamine and trimethyl-
amine, are charged at circumneutral pH. Precursors may
therefore have significantly different properties than
NDMA. As a result of their protonation, precursors
should be even less susceptible to treatment by air strip-
ping or adsorption compared to NDMA. Hwang et al.
(1994) found that dimethylamine and other aliphatic
amines were removed poorly by sorption to granular ac-
tivated carbon (Freundlich isotherm constants for di-
methylamine onto Calgon F-400 GAC were K 57.73
mg/g and 1/n50.26).
Although direct photolysis of NDMA is an effective
treatment technique, the lack of the nitroso functional
group on the nitrogen-containing precursors may make
precursors unreactive. Hwang et al. (1994) found that the
reaction of dimethylamine with ozone was slow. How-
ever, there are preliminary indications that hydroxyl rad-
icals formed in UV–hydrogen peroxide or ozone–hydro-
gen peroxide systems can remove NDMA precursors
NDMA precursors are readily removed by biological
treatment. Secondary biological treatment of municipal
398 MITCH ET AL.
wastewater was found to reduce NDMA precursors in
wastewater by an average of 60%, and to reduce di-
methylamine concentrations by at least an order of mag-
nitude (Mitch and Sedlak, 2002c). Although biological
nitrification and denitrification reduced NDMA forma-
tion during application of hypochlorite by precluding
monochloramine formation, these extended biological
treatments were not found to significantly reduce organic
NDMA precursor concentrations (Mitch and Sedlak,
2002c). Typical secondary treatment systems appear ca-
pable of removing the majority of biodegradable precur-
NDMA precursors also are removed in advanced treat-
ment systems. Microfiltration reduces the concentration
of particle-associated NDMA precursors in activated
sludge wastewater effluent (Mitch and Sedlak, 2002c).
Reverse osmosis treatment reduces NDMA precursor
concentrations by at least an order of magnitude, re-
moving not only colloidal NDMA precursors, but also
charged, dissolved precursors such as protonated di-
AREAS OF FUTURE RESEARCH
As indicated by this review, a substantial amount of
research has been performed on the source, behavior, and
treatment of NDMA. However, additional research needs
to be performed to develop more cost-effective means of
minimizing NDMA exposure. Several research needs are
Characterization of other N-nitroso compounds and
other products formed from the reaction of organic
nitrogen and monochloramine should be performed.
Characterization of the precursors responsible for
NDMA formation during chlorination of drinking
water should be elucidated better. In particular, ad-
ditional research is needed to determine the relative
importance of NDMA formation during passage of
water through ion exchange units or during post-
treatment chlorination of precursors leached from
ion exchange units relative to precursors from other
Characterization of organic nitrogen-containing
NDMA precursors during wastewater treatment is
needed. Only about 10% of the formation of NDMA
from organic nitrogen precursors in the low molec-
ular weight fraction of secondary wastewater efflu-
ent can be accounted for by dimethylamine. Other
dissolved precursors in wastewater must be identi-
fied. In addition, the role of treatment polymers as
potential NDMA precursors should be explored.
Methods are needed to improve the removal of or-
ganic nitrogen-containing NDMA precursors prior to
NDMA formation did not occur during ozone disin-
fection (Najm and Trussell, 2001). However, an in-
vestigation regarding the effect of ozonation on
NDMA precursors should be undertaken because
monochloramine may be applied to maintain a dis-
infection residual following ozonation. Furthermore,
the ability of alternative disinfectants such as chlo-
rine dioxide to form NDMA should be investigated.
Measurement of the quantum yield for aqueous
NDMA photolysis by sunlight is needed to predict
the rate of removal of NDMA upon sunlight expo-
sure in infiltration basins, during irrigation or in sur-
The uptake of NDMA by plants and its subsequent
fate should be further evaluated to determine the po-
tential for phytoremediation.
The pathway for microbial degradation of NDMA
and the associated kinetics should be evaluated to
identify conditions conducive to in situ bioremedi-
ation. The potential for NDMA removal in biofil-
tration treatment units should be evaluated as a
treatment strategy in drinking water treatment
We thank Mr. Timothy Durbin for help with this man-
uscript. Support for Mr. Bill Mitch was provided by the
WateReuse Association and the National Water Research
Institute. Support for Mr. Jonathan Sharp was provided
by funding from the University of California Toxics Sub-
stances Research and Teaching Program and by NIEHS
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