The effect of sulfamethoxasone on the growth, denitrifying activity, and community composition of groundwater bacteria from Cape Cod, MA

Material Information

The effect of sulfamethoxasone on the growth, denitrifying activity, and community composition of groundwater bacteria from Cape Cod, MA
Underwood, Jennifer Corinne
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75 leaves : ; 28 cm


Subjects / Keywords:
Sulfamethoxazole -- Massachusetts -- Cape Cod ( lcsh )
Bacterial pollution of water -- Massachusetts -- Cape Cod ( lcsh )
Groundwater -- Microbiology -- Massachusetts -- Cape Cod ( lcsh )
Denitrifying bacteria -- Massachusetts -- Cape Cod ( lcsh )
Bacterial pollution of water ( fast )
Denitrifying bacteria ( fast )
Groundwater -- Microbiology ( fast )
Sulfamethoxazole ( fast )
Massachusetts -- Cape Cod ( fast )
bibliography ( marcgt )
theses ( marcgt )
non-fiction ( marcgt )


Includes bibliographical references (leaves 66-75).
General Note:
Department of Integrative Biology
Statement of Responsibility:
by Jennifer Corinne Underwood.

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Source Institution:
|University of Colorado Denver
Holding Location:
|Auraria Library
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All applicable rights reserved by the source institution and holding location.
Resource Identifier:
516028004 ( OCLC )
LD1193.L45 2009m U52 ( lcc )

Full Text
Jennifer Corinne Underwood
B.A., Missouri State University, 2001
A thesis submitted to the
University of Colorado Denver
in partial fulfillment
of the requirements for the degree of
Masters of Science

This thesis for the Masters of Science
degree by
Jennifer Corinne Underwood
has been approved

The effects of wastewater-reflective concentrations of the antimicrobial
agent sulfamethoxazole (SMX) on the growth, denitrification activity, and community
structure of groundwater bacteria were assessed under laboratory conditions.
When groundwater bacteria isolated from a pristine zone of a sandy, drinking-water
aquifer on Cape Cod, MA were exposed to 1 pM (253.28 pg/L) SMX in the laboratory,
marked decreases in cell division were observed, as indicated by increased lag
phases and a 39.37% reduction in growth rates. The extent of inhibition tended to
increase within increasing SMX concentrations and the control was significantly (p <
6.36 x 10~4) different from SMX treatments (1 through 50 pM SMX). The addition of >
0.005 pM (1.3 pg/L), an environmentally-reflective concentration, resulted in
increased lag phases with one exponential phase compared to the control (0 pM
SMX) which appeared to have diauxic growth. Reduced denitrification activity
coincided with reduced bacterial growth rates as indicated by reduced N03'
reduction and N02' production rates upon exposure to 1 pM SMX (p < 0.03).
Exposure to 1 pM SMX led to 46% and 38% reductions in N03" reduction and N02
production rates in comparison to the control respectively. However, when the
numbers of bacteria were accounted for in rate calculations (pmol N03~or N02"

/day/cell) no trend on a per cell basis was observed. In contrast, denitrification
potential was reduced by exposure to > 0.005 pM SMX (p < 0.09) as indicated by
decreased quantities of N02 produced over the duration of the assay. The presence
of 0.004 pM (1 pg/L) and 0.04 pM (10 pg/L) SMX appeared to alter bacterial
community assemblages of groundwater bacteria based off results from a restriction
fragment length polymorphism. The presence of SMX appeared to shift the banding
pattern resulting in the appearance of new bands and the decrease or increase in
band intensities. Results of this study collectively suggest that environmentally-
reflective concentrations of SMX may alter groundwater bacterial cell growth and
reduce denitrification potential even though overall cell growth and denitrification
rates were not impaired until exposure to 1 pM SMX. In addition, the
environmentally-reflective concentrations of SMX may alter the composition of
bacterial communities. Thus, current concentrations of SMX in wastewater could
affect resident microbial communities and ecological function in aquifers.

This abstract accurately represents the content of the candidate's thesis. I
recommend its publication.

Thanks to my advisor, Timberly Roane, and my committee members, Ronald Harvey
and Lisa Johansen, for their support and guidance in my research. In addition, much
thanks to Deb Repert, Dave Metge, Dick Smith, Laura Baumgartner, and Norm Pace
for their guidance and their instrumental assistance with the technical aspects of
this project. Finally, I would like to thank Kate Holm and Kristen Peterson for their
assistance on this project.

1. INTRODUCTION.........................................................10
1.1 Background...........................................................10
1.2 Scope of the Thesis..................................................13
2. REVIEW OF THE LITERATURE................................................14
2.1 SMX and the Environment..............................................14
2.1.1 Definition, background and classification........................14
2.1.2 Application and consumption......................................16
2.1.3 Physiochemical properties and transport potential...............17
2.1.4 Fate and occurrence..............................................18
2.1.5 Degradation......................................................21
2.2 Ecotoxicology of sulfamethoxazole....................................23
2.2.1 Cell Division....................................................23
2.2.2 Microbial activity and denitrification...........................27
2.2.3 Species richness and community profile...........................31
3. STUDY LOCATION.......................................................33
4. METHODS..............................................................35
4.1 Groundwater Collection...............................................35
4.2 Laboratory Experiments...............................................36
4.2.1 Growth and Denitrification of Cultured Bacteria..................36
4.2.2 Growth and Community Composition.................................38
4.3 Analytical Techniques................................................39
4.3.1 Bacterial Stains and Epifluoresence Microscopy...................39
4.3.2 N03/N02 levels...................................................40
4.3.3 DNA Extraction...................................................40
4.3.4 PCR, Restriction Fragment Length Polymorphism and Cloning........41
5. RESULTS AND CONCLUSIONS..............................................42
5.1 SMX EFFECT ON BACTERIAL COMMUNITY GROWTH.............................42
5.2 SMX affect on Denitrification........................................50
5.3 SMX Affect on Community Composition..................................58
5.3.1 Restriction Fragment Length Polymorphism (RFLP)..................58

6. CONCLUSIONS..........................................................63
Analytical Methods..................................................74

Figure 2.1 Schematic representing the affect of SULs on the production of folic acid
in bacteria..........................................................15
Figure 3.1 Map of Cape Cod, MA and the location of the pristin (F605) and the
wastewater-impacted (F411) groundwater wells.........................34
Figure 5.1 Comparison of bacterial growth over 13 days upon exposure to 0, 0.005,
0.01, and 0.1 pM SMX.................................................45
Figure 5.2 Comparison of bacterial growth over 13 days upon exposure to 0,1, 5,
10, and 50 pM SMX....................................................46
Figure 5.3 Comparison of bacterial growth over 50 hours upon exposure to 0,
0.005, 0.01, and 0.1 pM SMX..........................................47
Figure 5.4 Differences in growth rates for SMX concentrations 0 through 50 pM... 48
Figure 5.5 Comparison of N03- N02- breakthrough curves for samples 0 through 50
pM SMX...............................................................52
Figure 5.7 Comparison of initial and final I\I03" and final ISI02" values for SMX
treatments 0 through 50 pM...........................................54
Figure 5.8 Comparison of final pmol N03 and N02 normalized to initial pmol N03'
for SMX treatments 0 through 50 pM...................................55
Figure 5.9 RFLP comparisons of SMX treatments 0, 0.004, and 0.04 pM and
untreated F411 and F605 bacterial communities using Mspl and HinPII.60
Figure 5.10 RFLP comparisons of SMX treatments 0.04 pM R3 and untreated F411
and F605 bacterial communities........................................61
Table 2.1 Physiochemical properties of SMX..................................18
Table 3.1 Phsyiochemical properties for sampled groundwater wells F605 and
Table 5.1 Comparison of N02' production rates for samples 0 through 50 pM SMX
with and without the use of bacterial counts for rate normalization......56

1. Introduction
1.1 Background
Growing human populations have led to an increasing need for drinking
water. Thus some densely populated regions are already using significant amounts
of raw water originating from wastewater effluent for drinking-water production
(Barnes et al. 2008; Bruchet and Mandra 1994; Levine and Asano 2004; Postel 2000).
Wastewater effluent can enter the drinking water supply via direct discharge into
surface waters or via ground infiltration to receiving aquifers. These wastewater-
impacted waters later become source waters for drinking water production for
nearby communities further downstream.
Within the U.S., groundwater supplies about 40% of the U.S. public water
supply and is additionally used by more than 40 million people who supply their own
drinking water via domestic wells. Groundwater represents the U.S. principal
reserve of available freshwater (75% polar ice and glaciers, 25% groundwater, and <
1% rivers, lakes, and soil moisture) (Barnes et al. 2008) and withdrawals have
increased in recent years (14% between 1985 and 2000 to 83.3 Bgal/day) (Hutson et
al. 2004). In addition to domestic use, groundwater is a major source of water used

for irrigation (Alley et al. 1999) and contributes to flow in many rivers and streams
(Barnes et al. 2008).
Recent surveys of U.S. water illustrate that valuable sources of freshwater
are vulnerable to contamination from a variety of organic wastewater contaminants
(OWCs). A United States Geological Survey (USGS) of groundwater reported the
detection of OWCs in 81% of groundwater sites (n=47) that were susceptible to
contamination from animal (or) human waste (Barnes et al. 2008). The USGS survey
of untreated drinking water sources (groundwater n= 20, surface water n= 49)
reported the detection of OWCs in all but 4 sites (Focazio et al. 2008). Thus, the
quality of wastewater effluent and source waters, such as groundwater, has come
under increased scrutiny due to the unknown consequences of OWCs to ecosystem
and human health.
The presence of antimicrobials in domestic waters has received special
attention because antimicrobial contaminants are produced with the specific
intention of killing or inhibiting bacteria. This has raised the concern about potential
negative impacts on environmental bacteria possibly leading to ecosystem
impairment as well as elevated risk to human health. Sulfamethoxazole (SMX) was
the most prevalent antimicrobial contaminant detected in groundwater according to

the 2000 USGS survey of United States groundwater (Barnes et al. 2008). SMX was
most frequently (23%, n= 47) detected with a peak concentration of 0.004 pM (1.11
pg/L) (Barnes et al. 2008). Additionally, the 1999-2000 USGS survey of potentially
contaminated U.S. streams frequently (19%, n= 139) detected SMX with a peak
concentration of 0.008 pM (1.9 pg/L) (Koplin et al. 2002). SMX has been measured in
the effluent of several wastewater treatment facilities (conventional activated
sludge or fixed-bed bioreactors) ranging from around 0.001 pM (0.3 pg/L) (Brown
et al. 2006; Gobel et al. 2007) to 0.024 pM (6 pg/L) (Batt et al. 2006). SMX was also
measured in the effluent of a high school septic tank was reported as 0.115 pM (29
pg/L) (Godfrey et al. 2007).
Due to the ubiquity of SMX in fresh water sources, there is concern about
potential consequences to environmental bacteria and, by extension, ecosystem
functioning. There have only been a few studies that have examined the
ecotoxicological impact ofSMx, or other sulfonamides, on environmental bacteria in
terms of standing stock (Al-Ahmad et al. 1999; Demoling et al. 2009; Thiele-Bruhn
and Beck 2005; Zielezny et al. 2006), soil activities (respiration, Fe II reduction, and
nitrogen processing) (Kotzerke et al. 2008; Thiele-Bruhn and Beck 2005; Vaclavik et
al. 2004; Zielezny et al. 2006), and community structure (Demoling et al. 2009;

Hammesfahr et al. 2008; Heuer et al. 2008; Isidori et al. 2005; Schmitt et al. 2005;
Zielezny et al. 2006). However, previous studies have concentrated on the effects of
sulfonamides on soil microbiology and there have not been any studies investigating
the ecotoxicological effect of SMX on groundwater bacteria. In addition, previous
studies have not implicated effects from SMX at the low concentrations typically
found in aquatic environments (0.001 to 0.1 pM). Because SMX readily
contaminates groundwater elucidation of the unknown consequences to bacterial
growth, ecosystem function, and bacterial community composition using
environmentally-reflective SMX concentrations in groundwater systems are critical.
1.2 Scope of the Thesis
The aim of this study was to investigate if, and at what concentrations, the
antimicrobial SMX has a negative impact on the growth, denitrifying activity, and
community structure of groundwater bacterial populations previously unexposed to
SMX. To address these aims two laboratory assays were employed.
Experiment 1: Exposure of cultured groundwater bacteria from a pristine
aquifer in Cape Cod, MAto increasing concentrations of SMX. Resulting bacterial
growth and denitrification activity was investigated using epifluorescence

microscopy for cell enumeration and ion chromatography to measure nitrate and
nitrite concentrations.
Experiment 2: Exposure of non-cultured groundwater bacteria from the
pristine aquifer to sterilized groundwater from the same pristine aquifer with
increasing concentrations of SMX. The differences in resulting bacterial communities
were investigated using restriction fragment length polymorphisms (RFLP) of 16s
rDNA PCR products. The resulting bacterial communities selected for using
environmentally-reflective concentrations of SMX were compared to the bacterial
community from a SMx-impacted groundwater source in Cape Cod, MA.
2. Review of the Literature
2.1 SMX and the Environment
2.1.1 Definition, background and classification
SMX, which is part of the drug class sulfonamides (SULs), is a synthetic
antimicrobial agent. SULs are the oldest class of antimicrobial drugs first used in
1935 (Stokstad and Jukes 1987). Antimicrobial SULs are bacteriostatic and exhibit a
broad spectrum of antimicrobial activity against Gram-negative and Gram-positive
bacteria as well as some protozoa by interfering with folic acid production. SULs

inhibit folic acid production in bacteria by acting as a structural analog and
competitive antagonist with p-aminobenzoic acid (PABA) for the active site of
dihydropteroate synthetase (DHPS), an enzyme that catalizes an essential reaction in
the synthetic pathway of tetrahydrofloric acid (THFA) (Fig. 2.1).
dihydropteroate diphosphate + p-aminobenzoic acid (PABA)
dihydropteroic acid
dihydrofofic acid
tetrahydnofolic acid
Figure 2.1 Schematic representing the affect of SULs on the production of folic
acid in bacteria.
SULs are competitive inhibitors to DHPS enzymes and inhibit incorporation of PABA
which is needed for dihydropteroic acid production (Martindale and Reynolds 1993).
SMX is often co-prescribed with trimethoprim and these act synergistically to
prevent folic acid production.

2.1.2 Application and consumption
SMX (brand name Gantanol) is used to treat urinary tract infections, acute otitis
media, trachoma, conjunctivitis, toxoplasmosis, malaria, and as a meningococcal
meningitis prophlaxis. It is also prescribed in conjunction with trimethoprim under
the brand name Bactrim. Although SMX has been used in veterinary medicine and is
apparently widely available for purchase with a veterinary prescription (Sarmah et
al. 2006) it is not approved by the Food and Drug Administration for therapeutic or
sub-therapeutic use in livestock in US (the FDA Approved Animal Drug List, the
Green Book
m042847.htm], 2008). However, other SULs s are approved (Council 1999) but only
account for 2.3% of the total amount of antimicrobials used for veterinary purposes
(Sarmah et al. 2006). Thus, human use appears to be the primary source of SMX
contamination. Although the exact estimates of total SMX usage are unknown
(Sarmah et al. 2006), approximately 23,000 tons of antimicrobials are produced
annually in the U.S. (USC 2001). Records of specific antimicrobial sales are not
tracked in the U.S. but in 2001, 53.6 tons of SMX was sold in Germany (Radke et al.

2.1.3 Physiochemical properties and transport potential
The propensity for SMX to enter aqueous environments such as groundwater
is largely due to the physiochemical properties of SMX (Table 2.1). SMX is highly
soluble at 610 mg/L (Barber et al. 2009), has a low sorption coefficients (Kd) of 37.6
adsorption/43.1 desorption (high organic content soil, pH 6.8) and 0.23
adsorption/0.40 desorption (low organic content soil, pH 4.4) (Drillia et al. 2005),
and a low K0w of 0.9 (Barber et al. 2009), all of which confer high mobility in the
At environmentally relevant pH, SMxcan be present as neutral (1.8 or as an anionic species (pH>5.7) because of its dissociation constants pKa (pKai: 2.0
and pKa2: 7.24) (Radke et al. 2009). The sorption of sulfonamides to organic and
inorganic materials can also be pH dependent resulting in the lowest sorption when
they are present as anionic species (Kahle and Stamm 2007; Radke et al. 2009).
Thus, under decreased sorption SMX has a high propensity to contaminate water

Table 2.1. Physicochemical properties of SMX
Molecular Weight 253.28 g/mol
Water solubility 610 mg/L (Barber et al. 2009)
Dissociation constant (pKa) pKa 1: 2 pKa 2: 7.24 (Joss et al. 2005)
Sorption coefficient (Kd) 37.6/43.1 (pH 4.3) and 0.23/0.40 (pH 6.8) (Drillia et al. 2005)
Partition coefficient (log Kow) 0.9 (Barber et al. 2009)
2.1.4 Fate and occurrence
After the incomplete metabolism of SMX within the human body the parent
compound and its metabolites are excreted via urine. Within 24 hours, 45-70% of
an oral dose of SMxis excreted but most is present in a partly metabolized form
(McEvoy et al. 2004). Of the total amount of SMxthat is excreted, 15-25% is present
as the unaltered parent form, 43% is present as N4-acetyl-SMx (Ac-SMX), and 9-15%
is present as SMx-Ni-glucuronide (SMX-Glu) with an additional 3 metabolites making
up the remainder (Gill et al. 1996; Koopmans and Van Der Meer 1995; Van der Ven
et al. 1994). However, once in the environment SMX-Glu and Ac-SMX rapidly
retransform (within 4 days) into SMX (Radke et al. 2009), and Ac-SMX has been

detected in equal concentrations to SMX in wastewater treatment plant effluent
(Gobel et al. 2004).
Once SMX and its potential metabolites are excreted they reach groundwater
through a variety of pathways that include direct discharge from wastewater
treatment plants, landfills, and application of animal and human waste to farmland
(Barnes et al. 2008). Due to its mobility, SMX can then percolate into aquifers and
the rate of permeation can be enhanced by rain events or anthropogenic methods
of groundwater recharge such as bank filtration.
It has been estimated that the average concentration of SMX in wastewater
influent in the U.S. is 0.013 pM (3.2 pg/L) (Huang et al. 2001) but SMX has been
detected in raw wastewater influent as high as 0.15 pM (38 pg/L) (Bhandari et al.
2008), wastewater effluent as high as 0.079 pM (20 pg/L) (Bhandari et al. 2008), and
was detected in the septic tank effluent from a high schools at 0.115 pM (29 pg/L)
(Godfrey et al. 2007).
Researchers have also investigated the occurrence of SMX in the
environment. A USGS survey of groundwater subject to potential exposure to
contamination from animal (or) human waste showed that SMX was the most
frequently detected antimicrobial with a peak concentration of 0.004 pM (1.11

pg/L) (Barnes et al. 2008). Another study of pharmaceutical contamination of
groundwater in the U.S. (Lindsey et al. 2001) reported similar SMX concentrations.
In a study by Standley et al. (2008) surface water ponds receiving discharge from on-
site septic systems through contaminated groundwater in Cape Cod, MA, were
reported to be contaminated by SMX, demonstrating the risk of contamination to
surface water supplied by groundwater flow.
In surface waters, SMX has been detected in similar concentrations to
contaminated groundwater (maximum detected concentration 0.007 pM or 1.9
pg/L) (Barber et al. 2006; Batt et al. 2006; Kolpin et al. 2002; Yang et al. 2004). SMX
has also been detected in untreated drinking source water (Focazio et al. 2008;
Stackelberg et al. 2007) but not detected in treated drinking water. A study by
Stackelberg et al. (2007) demonstrated that SMX was unable to persist after
conventional drinking water treatment due to moderate removal in the clarification
process (perhaps by ferric chloride coagulation resulting in acid or base hydrolysis)
and becomes undetectable following disinfection (due to oxidation with free
chlorine) (Stackelberg et al. 2007).

2.1.5 Degradation
The degradation of SMX in aquatic systems can occur through abiotic and
biotic degradation. Abiotic degradation includes processes such as photolysis,
hydrolysis, oxidation and reduction. Biodgradation is the elimination of the parent
compound by bacteria, fungi, or plants. However, it should be noted that
degradation does not necessarily refer to the complete degradation to its most basic
elements, a process referred to as mineralization. Partial degradation can result in
the accumulation of benign by products or by products even more toxic than the
parent compound. One recent example reported by Gomez et al. (2008) was the
increased toxicity to Daphnia magna from 4-methylaminoantipyrine, a dipyrone
hydrolyses product produced by application of photolysis. In addition, the
production of by products via bio- or photodegradation processes depend on
conditions such as temperature, composition of matrix, latitude, etc (Kummerer
As many variables affect biodegradation, the rate of biodegradation reported
in the literature is mixed. The rate of SMX degradation has been studied under low
dissolved organic carbon (DOC) concentrations typical of natural waters, such as
groundwater, and under high DOC concentrations encountered in wastewater

treatment. Under oxic conditions with low overall bacterial concentrations
(between 1000 10,000 cells/mL inoculums) and low DOC concentrations SMX
showed very low biodegradation rates (Al-Ahmad et al. 1999; Alexy et al. 2004). A
study by Al-Ahmad et al. (1999) showed 0% biodegradation after 40 days in a closed
bottle test using bacteria from wastewater effluent as an inoculum. A study by
Alexy et al. (2004) showed 4% biodegradation after 28 days in a closed bottle test
but showed 38% biodegradation when sodium acetate was added, possibly due to
co-metabolism. Under high bacterial concentrations and DOC concentrations like
those similar to wastewater treatment facilities the biodegradation of SMX was
substantial with biodegradation exceeding 90% in some cases (Batt et al. 2007; Drilla
et al. 2005; Gobel et al. 2007; Perez et al. 2005). A study by Drilla et al. (2005)
reported activation of SMX degradation when carbon is limited and ammonium is in
excess (potential situation in extended aeration systems), allowing SMxto act as
both a carbon or nitrogen source.
In surface waters, abiotic degradation of SMX may include hydrolysis or
photolysis. However, pharmaceutical manufacturers produce antimicrobials to be
resistant to hydrolysis signifying that direct or indirect photolysis is the primary
mechanism for abiotic degradation in surface waters (Andreozzi et al. 2003). As the

efficacy of photolysis can be influenced by a variety of environmental factors,
previous reports are mixed. Lam et al. (2004) reported that direct photolysis (direct
absorption of solar light) is important in limiting the persistence of SMX in sunlit
surface water {T1/2 = 2 hours in pure H20) using a sunlight simulator (Xe lamp as UV
source) but the presence of NO3', DOM, and bicarbonate, common in natural water,
decreased degradation rates. Anderozzi et al. (2003) reported a ti/2of 2.4 days using
a Hg lamp to simulate sunlight. Photodegradation efficacy was reduced with the
presence of N03', however the presence of humic mater acted as a photosensitizer.
Although these studies suggest that SMX is degradable by photolysis in
surface waters and biodegradation in wastewaters, SMX seems to resist natural
attenuation in groundwater. Barber et al. (2009) reported that SMX has persisted in
groundwater at Cape Cod, MA, for decades and over kilometers of transport.
Therefore, the aforementioned mechanisms of SMX degradation may not be
relevant if SMX reaches groundwater environments.
2.2 Ecotoxicology of sulfamethoxazole
2.2.1 Cell Division
SMX is synthesized with the intent to inhibit the microbial biochemical
pathway involved in the formation of folic acids. Folic acids are essential for

nucleotide development, thus are extremely important during cellular division.
Investigating the concentrations of SMX affecting growth of environmental bacterial
populations is important given that microorganisms are responsible for degradation
of numerous environmental contaminants and that microorganisms and microbial
activities are the basis of ecosystem health. Groundwater ecosytems are unique
because microorganisms make up the entire ecosystem as no higher trophic level
organisms survive.
Although clinical reports of minimal inhibitory concentrations (MIC) of
common bacterial pathogens range from 0.99 pM (0.25 mg/L) in Neisseria spp. to
505 pM (128 mg/L) in Enterobacteriacea (Andrews 2001) it is critical to understand
how SMX may impact environmental bacteria differently. Several studies have
investigated differences in bacterial growth when exposed to SMX or other SULs;
however most have tested the effects of SMX specifically on soil microorganisms.
SULs have been detected in soil systems between 1 and 11 pg/kg (Kemper 2008).
A study byTheile et al. (2005) investigated the effect of sulfapyridine on soil
bacterial growth in a sandy Cambisol soil by using fumigation-extracted microbial
carbon (Ec) as a growth measure. Ec represents the amount of organic carbon
extracted from chloroform fumigated samples minus the amount of organic carbon

extracted from non-fumigated samples. Their study found that spiking soil with
sulfapyridine led to a reduction of Ec. Over a 14 day incubation using 4011 pM
(1000 mg/kg) of sulfaphyridine plus a nutrient substrate (2000 mg C/kg, 90% milled
maize straw and 10% glucose) the Ec declined from 1190 to 450 mg/kg while fungal
biomass, which was analyzed by ergosterol (a sterol component of fungal cell
membranes), increased from 0.47 to 0.96 mg/kg. Sulfapyridine was concluded to
significantly (p < 0.05) reduce numbers of soil bacteria, resulting in a shift to a more
fungal dominated community.
A study by Demoling et al. (2009) investigated the effect of SMX on soil
bacterial growth using leucine incorporation. This type of analysis involved the
addition of H3-leucine as a substrate followed by visualization of H3-leucine
incorporation through radioautography. Test soils were either unamended or
amended with manure or alfalfa to enhance bacterial growth. A decrease in
bacterial growth was observed upon exposure to increasing SMX concentrations
(beginning with 79 pM or 20 mg SMx/kg soil) when soils were nutrient-amended, but
differences were not observed until exposure to 1974 pM SMX (500 mg SMx/kg soil)
in unamended soils. The growth stimulation provided by the addition manure or

alfalfa resulted in greater differences in growth rates when exposed to SMX because
SMX is a bacteriostatic antimicrobial.
A study by Al-Ahmad et al. (1999) conducted a growth inhibition test using
Pseudomonas putida, a Gram-negative environmental bacteria, and reported a 50 %
inhibition concentration (IC50) of 1 pM (256 pg/L) SMX which is reflective of the MIC
of susceptible pathogens.
Another SMX toxicity study used the bacterium Vibrio fisheri (marine
bacteria) to elucidate the acute toxicity of SMX while higher trophic level organisms,
eg. Ceriodaphnia dubia (crustacean cladocera), Pseudokirchneriella subcapitata
(green algae), and Brachionus calyciflorus (rotifer) were used to study chronic
toxicity (Isidori et al. 2005). The effective concentration for inhibition of 50 % of V.
fisheri (EC 5o) was 92 pM (23.3 mg/L), while the most sensitive species tested for
chronic toxicity was Ceriodaphnia dubia (crustacean cladocera) in which a lower EC50
of 0.83 pM (0.21 mg/L) was reported after 7 days.
Since SMX has been shown to readily contaminant surface and ground-water
sources the existing literature provides little information in regards to the effect SMX
has on bacterial growth in vulnerable ecosystems.

2.2.2 Microbial activity and denitrification
Understanding how SMX may influence bacterial metabolic activities is vital
given the importance of bacteria in biogeochemical processes in the environment.
Several studies have examined the effects of SULs on microbial activities, such as soil
respiration, and in each case carbon substrates were required in order to observe
effects. Because SMX impacts DNA replication and cell division, carbon substrates
were needed to stimulate growth making the cells susceptible to SMX exposure. In
addition, for SULs to be toxic, active transport of SMX into the cell is necessary (Grafe
1992). On the other hand, the majority of soil microorganisms are dormant
(Jenkinson and Ladd 1981) and maintain respiration by oxidizing internal energy
sources (Tate III 2000), reducing their susceptibility. It is also common for the
observed effects of SUL exposure to be delayed for several hours due to cellular
reserves of growth factors, such as folic acid, the main target of SMX toxicity (Halling-
Sprensen 2001).
Zielenzy et al. (2006) showed how sulfadiazine affected soil bacterial
respiration using a respirometer. Differences in respiration were observed when soil
was amended with 5 g glucose/kg soil, but no differences in respiration were
observed when soil was not provided with glucose. With the glucose amendment,

decreased oxygen consumption after 36 hours was reported with increasing
sulfadiazine concentrations (beginning with 4 pM or lmg/kg soil). It was determined
that dose-dependent reductions in oxygen consumption occurred, but only after
bacterial growth was stimulated.
Theile-Bruhn and Beck (2005) studied the effects of sulfapyridine on soil
microbial basal respiration, dehydrogenase activity, substrate-induced respiration
(SIR), and Fe (II) reduction. They found no differences between soils with or without
sulfapyridine, up to concentrations as high as 4011 pM (1000 mg/kg), using basal
respiration and dehydrogenase activity assays; however when glucose was added
pronounced differences were observed using SIR and Fe (II) reduction assays. The
lowest effective dose (ED) of sulfapyridine that affected 10% (EDi0) of bacteria using
the SIR assay in Camisol soils was 0.2 pM (0.05 mg/kg) while 50% (ED50) of bacteria
were affected by 24.9 pM (6.2 mg/kg). The Fe (II) reduction assay using Luvisol
soils resulted in an EDio and ED50 of 0.01 pM (0.003 mg/kg) and 25.9 pM (6.45
mg/kg) respectively. Vaclavik et al. (2004) used sulfachloropyridazin to test soil
respiration and reported an ED50 of 252.89 pM (72 mg/kg) and a dose-dependent
response beginning with 3.51 pM (1 mg/kg). Similar to the report by Zieleny et al.

(2006), the addition of substrates was a requirement to observe effects from a SUL
on respiration or Fe (II) reduction.
Denitrification, the dissimilatory process that reduces nitrate (NO3 ) and
ultimately produces nitrogen (N2) gas through a series of intermediates, is an
important groundwater process that has been well documented within the aquifer
at Cape Cod, MA (Repert et al. 2006; Smith et al. 2004; Smith et al. 1994; Smith and
Duff 1988; Smith et al. 2001). Pristine environments are, in large part, extremely
nitrogen limited, however, anthropogenic sources of N03, such as from wastewater
and agriculture, make N03 one of the leading forms of nutrient contamination in
groundwater. The federal primary drinking water standard for NO3-N is 10 mg/L.
However, NO3-N concentrations of < 0.2 mg N/L generally represent natural
conditions, whereas concentrations > 3 mg N/L generally indicate effects from
human activities. Even in comparably low concentrations, N03' can be hazardous
to the health of human infants, yet in parts of the U.S., more than 25% of
groundwater wells exceed 3 mg/L N03 -N (Spalding and Exner 1991). Denitrification
is such a significant process in the removal of groundwater NO3' that it has been
documented to remove 41.7 mega g N/yr within a 12 m-deep 15 km-long transect of
a contaminated aquifer (Spalding and Parrott 1994). In many circumstances,

denitrification is the only feasible and environmentally sound way to remove
groundwater N03" on a large scale (Soares 2000). However, in the U.S., biological
treatment has not been approved as a method of treatment for potable water
(Soares 2000). Due to the importance of natural denitrification processes as a way
to remove N03 contaminants from groundwater, it is imperative to understand the
effect that SMX, which may co-occur with N03" contamination from wastewater, has
on denitrification rates.
Only a few studies have examined the effect of antimicrobials on
denitrification (Colloff et al. 2008; Costanzo et al. 2005) and no studies are available
addressing SMX effects on denitrification in aqueous systems. In a study by Costanzo
et al. (2005), bacteria isolated from creek sediment downstream of a wastewater
treatment plant showed differences in denitrification rates using 1000 pg/L of
erythromycin, clarithromycin, and amoxicillin. A study by Collof et al. (2008) looked
at the presence genes relating to the nitrogen cycle (nitrification, denitrification, and
nitrogen fixation) in soil bacteria from contrasting soils that were exposed to a
mixture of antibiotics, which included 5 g/kg of SMx-penicillin (50:50). It was found
that high-level exposures led to the rapid disappearance of the amok gene
(nitrification gene) and coincided with a rapid accumulation of ammonia in all soils

tested. After SMX exposure, the napA gene (N03" reduction) was absent from sugar-
cane soil within 24 hours and progressively decreased in abundance over a period of
60 days in soils from Wilby, Victoria (grain, legume, and cereal rotation soil with
temperate climate) and Moora (cereal, legume rotation with Mediterranean
climate) (Colloff et al. 2008). However, this did not result in alterations of the N03'
concentration. It was concluded that denitrifying bacteria may be better able to
withstand the presence of antimicrobials compared to nitrifying bacteria.
2.2.3 Species richness and community profile
Several studies investigated how SULs affect bacterial diversity in soil
systems using community profiling. Demoling et al. (2009) investigated the effect
SMX had on a soil microbial community structure using phospholipids fatty acid
(PLFA) analysis and community-level physiological profiling (CLPP). PLFA was used to
determine phospholipid fatty acid types within the community, while CLPP helped
determine shifts in the organic substrate use of a community. In alfalfa-amended
soils, an increase in fungal PLFAs with a corresponding decrease in bacterial PLFAs
occurred upon exposure to 78.96 pM SMX (20 mg SMx/kg soil). This was contributed
to a decrease in bacterial growth rates. The use of CLPP was a less sensitive form of

measurement than PLFA, however a significant difference in carbon substrate use at
the highest concentrations of SMX (1975 pM SMX or 500 mg SMx/kg soil) was shown.
Both types of analyses provide information on pollution induced community
tolerance (PICT). An increase in PICT implies that those tolerant to a particular
contaminant will increase while those sensitive will decrease in abundance. PICT
was also investigated using CLPP by Schmitt et al. (2005) on soil spiked with
sulfachloropyridazine. An increase in PICT by 10% was found using 25.63 pM (7.3
mg sulfachloropyridazine/kg soil) and a decrease in metabolic diversity was only
observed at the highest concentration (3512 pM or 1000 mg sulfachlorpyridazine/kg
Using analyses of 16S rDNA, Zielenzy et al. (2009) found that in the presence
of glucose, the bacterial community structure was affected by sulfadiazine (SDZ) in
dose-dependent quantities. Using PCR-denaturing gradient gel electrophoresis shifts
in bacterial communities were observed. When comparing soils amended with
glucose only and glucose plus 1 mg or 10 mg SDZ/kg soil, community profiles
changed with increasing SDZ concentrations as indicated by the appearance of new
bands or the increase of band intensities. However when 50 mg SDZ/kg soil was
used, the band pattern was most similar to the control (without glucose and SDZ),

presumably due to the strong growth inhibition which is in accordance with
measured static respiration rates.
3. Study Location
The bacterial communities were collected from two well sites within an
unconfined sandy aquifer in Cape Cod, MA (Fig. 3.1). Groundwater well F605 is
located within the Crane Wildlife Management Area. This groundwater location has
not been in the path of any known contaminant plume and is considered pristine for
the purposes of this study. At well F605, groundwater was sampled from a depth of
10 meters below land surface. Groundwater well F411 is located within the core of
the Ashument Valley Wastewater Plume. This plume resulted from the
Massachusetts Military Reservation's use of infiltration beds for disposal of treated
wastewater. When use of the beds was discontinued in 1995, the existing plume
contained 0.006 pM (1.5 pg/L) SMX, 0.6 pg/L of the solvent tetrachloroethene, 0.25
pg/L of the disinfectant 1,4-dichlorobenzene, and 0.6 pg/L of the nonionic
surfactant degradation product 4-nonylphenol when sampled in September 2005
(Barber et al. 2009). At well F411, groundwater was sampled from a depth of 3.5
meters below land surface. Physicochemical properties for the aquifer at each well
location are provided in Table 3.1.

.. I '
Figure 3.1 Map of Cape Cod, MA and the location of the pristin (F605) and the
wastewater-impacted (F411) groundwater wells.
Illustration adapted from

Table 3.1. Physiochemical properties for sampled groundwater wells F605 and F411
Physiochemical properties F605 (pristine) F411 (wastewater-impacted)
pH 5.55 6.22
DOC 0.4 1.6
DO (mg/L) 9.96 0.27
Conductivity (pS/cm) 51.3 166.2
Temperature (C) 10.1 12.7
4. Methods
4.1 Groundwater Collection
Groundwater samples were taken from well F605 in March 2009 from a
depth of 10 meters with a multi-level sampler constructed of 1.25 inch PVC pipe
with 0.25 inch polypropylene tube placed inside with a nylon screen attached at the
base. 110 L of water was removed using a peristaltic pump. The stability of the
specific conductance was checked prior to sample collection. Bulk samples were
stored in two sterile 5 L carboys, packed on ice, and then shipped via overnight
delivery to the USGS laboratories in Boulder, Colorado, U.S.A.

Groundwater was sampled from well F411 in September 2009 from a depth
of 3.6 meters in the same manner as described for well F605. The bulk sample was
stored in one sterile 1 L amber bottle, packed on ice and then shipped overnight
delivery to the USGS laboratories in Boulder, Colorado, U.S.A.
4.2 Laboratory Experiments
4.2.1 Growth and Denitrification of Cultured Bacteria
In order to study differences in bacterial growth and denitrification rates
laboratory incubations were performed. Sterile serum bottles (150 mL) containing
100 mL of O.OOlx tryptic soy broth (TSB) amended with 2 mM NaN03 (in excess)
were loosely covered with foil and allowed to degas overnight in an anaerobic glove
box containing 95% l\l2 and 5% H2 gas mix. Dilute TSB was chosen because it is an
undefined, non selective, heterotrophic medium that we hypothesized would allow
for stability of a diverse microbial community. Serum bottles for bacterial growth or
denitrification analyses, in triplicate, were then capped using rubber stoppers and
aluminum crimpers and autoclaved for 20 minutes at 121C. After cooling to room
temperature, serial dilutions of SMX dissolved in anaerobic deionized water were
injected into each bottle using sterile needles, syringes, and 0.1 pm (pore-size)

syringe filters. The final SMX concentrations were: 0, 0.005, 0.01, 0.1,1, 5,10, 50,
100, 500,1000, and 2000 (pM). Data is only presented for concentrations 0 through
50 pM because the higher SMX concentrations showed no significant growth or
denitrification activity within the duration of the experiment. Finally, each serum
bottle was inoculated with 8 mLof the bacterial community from the pristine
groundwater (F605) grown in O.OOlxTSB amended with 2 mM NaN03 and
incubated at room temperature (21C) for 7 days. The initial titer for each serum
bottle was ~ 6.7xl03 cells/mL. All bottles were then rotated at 70 rpm at ~ 21C.
Serum bottles for bacterial growth and denitrification were periodically sampled
using sterile, helium-flushed syringes that were aseptically injected through the
rubber stoppers to withdraw liquid volumes from 1 to 10 mLs for cell growth
analyses depending on the cell density and 10 mLs for N03' and N02" analyses.
Bacterial growth samples were vacuum filtered using 0.22 pm membrane filters atop
0.8 pm backing filters, until ~ 1 mL of sample remained to be filtered at which point
Yi mL of lOmg/L of the nucleic acid stain 4',6-diamidino-2-phenylindole (DAPI) was
applied. The samples stained in the dark for ~ 20 minutes. After staining, the
remaining sample was filtered and the membrane filter was removed and placed on
a glass microscope slide. A drop of immersion oil was placed on the filter and then

covered with a cover slip. The prepared slide was placed in a slide box and stored in
the freezer at -209 C. Samples for denitrification analyses were filtered using a 0.2
pm pore-size syringe filter and the filtrate was frozen at 20 C.
4.2.2 Growth and Community Composition
In order to study how SMX may impact groundwater bacterial community
composition an additional growth experiment was conducted. Groundwater
collected from well F605 was heat sterilized using an autoclave at 121 C and 200 mL
aliquots were transferred to capped 250 mL polypropolene copolymer nalgene
bottles. After cooling to room temperature, serial dilutions of SMX, which had been
dissolved in sterile deionized water, was added to each container followed by the
addition of 1.5 mL of non-sterilized F605 groundwater. The final SMX concentrations
were: 0, 0.004, 0.04, 0.4, 4, 20, 40, 200, and 400 pM in triplicate. Data is only
presented for concentrations 0 through 0.04 pM because differences in community
composition were already apparent. The initial bacterial concentration was ~ 4 x 103
cells/mL. The bottles were placed on a shaker table at 70 rpm for 3 weeks and 3
days (2/13/09 to 3/06/09). Samples (45 mL) were collected and stored in sterile
conical vials at 20 C until further analysis. In addition, a 200 mL sample of

untreated F605 was immediately stored at 20 C and a 200 mL sample of
autoclave-sterilized F605 was stored at 20 C after 2 weeks of incubation.
4.3 Analytical Techniques
4.3.1 Bacterial Stains and Epifluoresence Microscopy
The nucleic acid stain, DAPI (catalog D3571, Invitrogen, Carlsbad, CA) was
used for total cell counts. An epifluorescent microscope was used to enumerate
cell density (#cells/mL) using 630x or lOOOx magnification. Ten fields of view were
counted on each filter that had a cell density of 30 100 cells/field. The following
formula was used to derive cell density:
# bacterial cells/mL = (average # cells/field area of view/field)/volume (mL)
Growth rates were produced via analyses of linear relationships at the exponential
phase of each growth curve. Growth rates are reported as p which is determined by
the following equation:
p = ln2/generation time (h

4.3.2 N03'/N02'levels
Samples to be assayed for N03' and N02" content assessed using a Dionex DX-
120 Ion Chromatographer using an Ion Pac AS514 column. This instrument has a
minimum detection limit of 0.2 pM (12.4 pg/L) for N03 and an unknown detection
limit for N02. An isocratic eluent of 4.3 mM Na2C031.2 mM NaHC03 was used with
a 1.25 mL/min flow rate. Sample concentrations were derived by comparison to a
prepared standard curve that was referenced to a USGS standard reference sample.
Duplicate measurements were taken on 10% of samples with percent error < 5%.
4.3.3 DNA Extraction
DIMA was extracted from thawed groundwater samples using a bead beating
protocol adapted from Dojka et al. 1998. The samples were vacuum filtered using
Nalgene (Rochester, NY) analytical test filter funnel (145-0020) onto sterile 25 mm-
diameter, 0.2 pm-pore Millipore (Billerica, MA) Isopore membrane filters GTTP
(GTTP 04700). The cells then were reciprocated on a Mini-Beadbeater 8 (Biospec
Bartlesville, OK) at high speed (2800 oscillations/min) for 2 min. upon the addition of
50% (vol/vol) phenolchloroform-isoamyl alcohol (24:24:1), 5% (wt/vol) sodium
dodecyl sulfate, and approximately 0.3 g of acid-washed zirconium-silica beads (0.1
mm diameter). Following cell lysis, DNA from 300 pL of lysate was precipitated via

25 min. of centrifugation at 10.6 rpm with the addition of 7.5 pL 0.7 % (vol/vol) 1 M
glycogen, 200 pL 14% 7.5 M ammonium acetate, plus the addition of an equal
volume (500 pL) of isopropanol. DNA was additionally precipitated via 5 min. of
centrifugation at 10.6 rpm using 70 % ethanol. Following supernatant removal, the
resulting DNA pellet was re-suspended in 50 pL of lx TE buffer.
4.3.4 PCR and Restriction Fragment Length Polymorphism
Community 16S rDNA was PCR amplified in reactions containing 1 pL DNA, 10
pL Hot MasterMix (5 prime Gaithersburg, MD), 10 pL DNA-free H20, 2 pL (lmg/mL)
BSA, and 1 pL (0.2 pM) of each forward and reverse primer. The primer pair were
forward 515F (GTGCCAGCMGCCGCGGTAA) and reverse 1391R
(GACGGGCGGTGTGTRCA). A model PT-100 thermal cycler (MJ Research Inc.) was
used for the amplification cycles: 94C for 2 min (for initial denaturation and
activation of AmpliTaq Gold), followed by 32 cycles at 94C for 20 s, 52C for 20 s,
and 65C for 1 min 30 s and a final extension period of 10 min at 65C. To confirm
PCR amplification, 5 pL of PCR product and 1 pL of 6x loading dye was run on a 1%
agarose gel in 1% TBE buffer at 120 V for 20 min.

Following amplification and confirmation via gel electrophoresis, RFLP
analysis was performed using 50 pL of PCR product that was incubated overnight at
37 C in the presence of 6 pL of lx NEB 2 (New England Biolabs, Natick, MA) plus
DTT, 8 pL of loading dye, and 0.4 pL each of restriction enzymes Mspl and HinPII. The
samples were then size separated on a 3 % agarose gel with 3 pL/lOOmL of ethidium
bromide at 120 V for 3 hours at 4 C and viewed at 312 nm.
5. Results and Conclusions
5.1 SMX effect on Bacterial Community Growth
Due to the direct relationship between microbial growth and metabolic rates
(Russell and Cook 1995; Monod 1949), it was important to determine how different
concentrations of SMX affected the growth of the bacterial community from the
pristine aquifer. It would be expected that if bacterial growth was impaired due to
the presence of an inhibitor such as SMX, then the oxidation (of TSB) and reduction
(of N03 ) rates would also be constrained. However, the level of SMX required to
inhibit growth first needed to be determined. There is limited research on the
ecotoxicological effects of SMX on environmental bacteria. The lowest reported
effect from SMX was an IC50 from 1 pM (256 pg/L) SMX (Al-Ahmad et al. 1999) using
P. putida, a much higher concentration than what has been observed in aquatic

environments. The least sensitive reported effect from the use of the SUL
sulfaphyridine was 4011 pM (Theile et al. 2005). The highest reported
environmental level of SMX (0.115 pM (29 pg/L)) was for septic tank effluent at a
high school (Godfrey et al. 2007), but concentrations in contaminated groundwater
have not been shown to exceed 0.006 pM (1.5 pg/L) (Barber et al. 2009). Therefore,
the growth response of bacteria from the pristine aquifer was assessed using a wide
range of SMX concentrations (0 through 2000 pM), including the environmentally
reflective concentrations of 0.005 and 0.01 pM. Surprisingly, effects were observed
at sub-therapeutic concentrations (a therapeutic dose of Bactrim is 800 mg SMX for
adults with urinary tract infection and 75 to 100 mg/kg or 296 to 395 pmol/Kg for
adults with Pneumocystis carinii pneumonia over 24 hours), and so 0 through 50 pM
SMX are reported here. Bacteria exposed to SMX concentrations of 100 pM or
greater did not grow during the 13 days of observation (data not shown).
There were noticeable differences in cell growth between bacteria exposed to
increasing concentrations of SMX. The control (0 pM SMX) contrasted from all SMX
treatments (0.005 through 50 pM) by not experiencing an observable initial lag
phase. Instead, the control had an initial rapid period of growth followed by a lag
phase then a second slower period of growth (Fig. 5.1, 5.2, 5.3). All SMX exposed

samples had an initial lag phase followed by one exponential phase and a stationary
phase. Ideally, two growth rates should be determined for the control representing
the two periods of growth. Instead, the growth rate for the control was averaged
for overall growth because additional data points would be necessary to provide
reliable rates for two different periods of growth (Fig. 5.3). Although exposure to
SM* increased lag times, there were no inhibitory effects on bacterial growth rates
compared to the control until the exposure of 1 pM SMX (Fig. 5.4). The exposure of
1 pM SMX led to a 39.37 0.96% reduction in growth rates compared to the control.
Previously measured growth rates for unattached bacteria within the contaminated
part of the aquifer ranged from p = 0.043 0.005 which was similar to the measured
growth of SMX treatments 1 through 50 pM (Harvey and George 1987) (Fig. 5.4).
Thus, this experiment allowed for a higher rate of growth than what is reflective of
the contaminated part of the aquifer. Samples also differed in the duration of the
time until start of stationary phase and maximum cell density (Fig. 5.1, 5.2).
Treatments 0, 0.005, and 0.01 pM reached stationary phase at ~ 55 hour, 0.1 pM at
~ 77 hours, and 1 pM at ~ 144 hours. SMX treatments 0 through 1 pM SMX reached
a maximum cell density of 3.6 x 106 1 x 106 cells/mL. Growth of SMX treatments 5
though 50 pM stopped at ~ 148 hours with a maximum cell density of 8.1 x 105 6.3

x 105 cells/mL for treatments 5 and 10 pM and 8.7 x 104 1.5 x 103 cells/mL for 50
0 |iM
0.005 pM
0.01 pM
0.1 pM
Figure 5.1 Comparison of bacterial growth over 13 days upon exposure to 0,0.005,
0.01, and 0.1 pM SMX.
Error bars represent standard error of n=3.

I' : i i
1E+06 s S 1 11
I -
= 1E+05 I i S q ro 0 i
U 00 ip 1 0 5
- 1E+04 c 3
1E+03 -- -- --- t
0 2 4 6 8 10 12 14
& lpM
10 (iM
50 pM
Figure 5.2 Comparison of bacterial growth over 13 days upon exposure to 0,1, 5,
10, and 50 pM SMX.
Error bars represent standard error of n=3.

> 3E+05
~ 3E+04
3E+03 -
1 0 |iM
0.005 pM
0.01 pM
0.1 pM
12 24 36 48 60
Time (hours)
Figure 5.3 Comparison of bacterial growth over 50 hours upon exposure to 0,
0.005, 0.01, and 0.1 pM SMX.
Error bars represent standard error of n=3

0 0.005 0.01 0.1 1 5 10 50
Figure 5.4 Differences in growth rates for SMX concentrations 0 through 50 pM.
Error bars represent standard error of n=3.
This study found that SMX exposure, at concentrations as low as 0.005 pM,
and reflective of environmental concentrations, could have negative consequences
on cellular division for some groundwater bacteria as indicated by increased lag
phases. However, it is unknown if the increased lag phase upon exposure to 0.005
pM SMX would have environmental consequences as growth rates were not
inhibited until exposure to 1 pM SMX, a concentration that is comparable to
previous reports of growth inhibition (Al-Ahmad et al. 1999; Andrews 2001). The
growth behavior of the control appeared to resemble a diauxic growth curve as

indicated by a period of rapid growth (p = 0.31 0.02 calculated off of 2 data points,
data not shown), a secondary lag phase, and a period of slower growth (p = 0.04
0.00 calculated off of 2 data points, data not shown). All SMx-treated samples
resulted in a standard growth curve with one exponential phase. It is unknown why
the growth behavior of the control contrasted from all SMx-treated samples or if it
has any environmentally-relevant consequences. The control could have had
differential growth rates due to a variety of reasons including a shift in the bacterial
community or changes in substrate consumption. However, the potential that the
differential growth rates could be as a result of an enzymatic shift in substrate use
from a sugar the bacteria could directly metabolize to a less preferential sugar,
characteristic of diauxic growth (Ullmann 2007), is intriguing due to potential
environmental relevance. If the SMx-treated communities were not able to utilize
the easily metabolized carbon source present in TSB, or potentially in the
environment, then they would be able to perform less work due to this limitation. In
order to elucidate some of these questions further research would be required that
included determination of bacterial communities and variable carbon concentrations
at different time points along bacterial growth curves.

5.2 SMX effect on denitrification
The exposure of SMX led to a trend of reduced denitrification rates that
corresponded with reduced growth rates (Fig. 5.5, 5.6). The exposure of 1 pM SMX
led to a 46 15.2% reduction in N03 removal rates and a 38 3.5% reduction in
N02" production rates compared to the control (Fig 5.5). This is similar to the 39.37
0.96% reduction in growth rates when exposed to 1 pM SMX. In addition, the onset
of N02" production experienced a lag time which was apparent in concentrations 1
through 50 pM SMX (Fig. 5.5,). Since each SMx-treatment type resulted in a different
concentration of bacteria throughout the experiment due to differences in cell
growth, it was important to determine if the relationship of reduced denitrification
rates with increased SMX concentrations still existed when accounting for the
number of bacteria in each sample at each time point. When denitrification rates
were normalized for total number of bacteria/sample no trend in denitrification
activity was observed (table 5.1). Thus, it appeared that the reduction of
denitrification rates was a product of reduced growth rates rather than a direct
impact on enzymatic activity. Flowever, it cannot be assumed that all the bacteria
present in each sample participated in denitrification. Thus, it cannot be concluded

that SMX exposure did not directly impact denitrification activity in some way by the
results of this experiment.
Conversely, the decreased amount of N03 removed and NO2" produced over
the 19 day incubation correlated with increasing SMX concentrations, beginning with
0.005 pM SMX (Fig. 5.5, 5.7, 5.8). The control produced more conversion to NC>2~(p <
0.09) than all SMx-treated samples (Fig. 5.7, 5.8). The lowest SMXtreatment (0.005
pM) produced 6.9 3.3% less NO2" than the control and 1 pM SMX produced 29.8
1.9% less. The use of NO2" appears to be a more accurate measure of denitrification
activity in this assay as compared to I\I03". This is likely due to difficulties detecting
differences in small changes of N03" concentrations while using high initial N03
concentrations (2.2 mM 0.16).

0.005 nM
0.01 |iM
ft ftft ft
u I 1 f* I I : K ** B

/.. ~ 52 ;
N 03-
10 20 0 10 20 0 10 20 0 10
1 nM 5 nM 10 (iM 50 nM
0 10 20 0
10 20 0 10 20 0 10 20
Figure 5.5 Comparison of N03- N02- breakthrough curves for samples 0 through
50 pM SMX.
Error bars represent standard error n=3, except 0.01 and 50 pM are n=2.

I Nitrate
Figure 5.6 Comparison of N03 removal and N02 production rates for samples 0
through 50 pM SMX. Error bars represent standard error n=3.

* r*j
' m
Initial Nitrate
. Final Nitrate
Final Nitrite
Figure 5.7 Comparison of initial and final N03 and final N02' values for SMX
treatments 0 through 50 pM.
The incubation lasted for 19 days. Error bars represent standard error n=3, except
0.01 and 50 pM are n=2.


Q>> \ <3 ^ I Nitrate
Figure 5.8 Comparison of final pmol N03' and N02' normalized to initial pmol N03
for SMX treatments 0 through 50 |iM.
The incubation lasted for 19 days. Error bars represent standard error n=3, except
0.01 and 50 pM are n=2.

Table 5.1 Rates of N03 removal and N02 production with and without control for
cell growth.
pM SMX pmol N03 /day S.E. jumol N02' /day S.E. fmol S.E. N03'/day /bacterium fmol N02'/day /bacterium S.E.
0 16.47 3.11 11.41 1.59 3.54 0.03 0.02 '* 0.00
0.005 20.08 3.84 14.74 0.16 16.41 0.18 0.02 0.00
0.01 16.86 0.31 12.78 0.55 2051 0.08 0.03 £ 0.00
0.1 16.44 2.21 9.90 0.41 26.16 0.58 0.02 0.00
1 6.14 0.10 5.11 0.34 4.64 0.03 0.00 0.00
5 1.82 0.69 1.37 0.43 10.35 1.41 '0.02 ^ 0.01
10 1.83 0.10 1.16 0.11 0.06 0.02 0.06 0.00
50 0.13 0.10 0.05 0.00 -4.42 0.21 o.o5 - 0.00
Few studies have examined the effect of antimicrobials on denitrification
activity (Colloff et al. 2008; Costanzo et al. 2005), especially at environmentally-
reflective concentrations. In a study by Costanzo et al. (2005), bacteria isolated from
creek sediment downstream of a wastewater treatment plant showed differences in
denitrification rates when exposed to 1000 pg/L of erythromycin, clarithromycin,
and amoxicillin. A study by Collof et al. (2008) looked at the expression of different
genes relating to the nitrogen cycle in soil bacteria from soils exposed to a mixture
of antibiotics, including (19.7 mM) 5 mg/g of SMx-penicillin (50:50). It was found
that such high-level exposures led to the rapid disappearance of the omoA gene
(nitrification gene) and coincided with a rapid accumulation of ammonia in all soils

tested. The nopA gene (N03' reduction) was absent from a sugar-cane soil within 24
hours after SMX exposure; however, this did not result in alterations of the expected
NO3' concentration (Colloff et al. 2008).
Conversely, this study indicated negative impacts on denitrification potential
(most NO3" reduced and N02" produced) at environmentally reflective
concentrations using groundwater bacteria. However, It is unclear why the control
had the highest denitrification potential over the duration of the experiment as
compared to SMx-treated samples while rates of N03" reduction and N02 production
were not affected until exposure to 1 pM SMX. Additionally, exposure to 1 pM SMX
seemed to be attributed to reduced growth rates rather than a direct effect upon
denitrification enzymatic activity. This was an expected result as previous studies
required a carbon amendment to stimulate growth and hence, observe differences
in soil activities between treatments.
Although it is unknown why a dose-dependent relationship between
increasing SMX concentrations and decreasing denitrification potentials was
observed, it may have to do with the differences between the observed growth
behavior between the control and the SMx-treated samples. If the growth behavior
of the control was diauxic as a result of an enzymatic shift in carbon substrate use,

perhaps the control produced the greatest amount of N02" because more carbon
sources were available for denitrification. This idea is feasible as a variety of carbon
sources were available but in limited quantities in this experiment (initial
concentration was 12.4 mg/L). This could have important human health and
ecological consequences as N03 is one of the leading groundwater contaminants in
the U.S. High levels of N03' can lead to methemoglobinemia in infants or
eutrophication in rivers and lakes. In groundwater, bacteria perform a dominant
role in the natural attenuation of N03~. If groundwater bacteria could only
metabolize select carbon sources as a result of SMX contamination, then their overall
denitrification potential could be reduced.
5.3 SMX Affect on Community Composition
5.3.1 Restriction Fragment Length Polymorphism (RFLP)
RFLPs of incubations using autoclave-sterilized F605 groundwater (no
additional amendments) with (0.004 and 0.04 pM Rl, R2, R3) or without (0 pM Rl,
R2, R3) the addition of SMX produced fewer bands compared to the RFLP profile
from F411 and untreated F605 bacterial communities (Fig. 5.8). The two negative
controls were autoclave-sterilized F605 groundwater and uncut DNA, which were

used to confirm that autoclaving kills cells and denatures DNA and to confirm that
PCR product was cut by restriction enzymes (Fig. 5.8). These community profiles,
generated following a 24 day incubation, compared the bacterial growth from a 1.5
mL inoculation of fresh F605 groundwater bacteria into autoclave-sterilized F505
groundwater without SMX or with 0.004 and 0.04 pM SMx. Thus, some bacteria
from F605 may not have been able to grow during this time period and/or under the
conditions provided resulting in a biased representation of the full bacterial
community from F605. This is reflected in the RFLP banding patterns when
comparing the control and the untreated F605 bacterial community (Fig. 5.8).
However, the purpose of this experiment was to determine if environmentally
reflective concentrations of SMX could alter the structure of a pristine groundwater
bacterial community, as represented by an alteration in resulting banding patterns
created by the use the restriction enzymes Mspl and HinPII. In addition, a
restriction digest of F411 (wastewater and SMx-impacted) bacteria was used to
compare to the banding pattern of F605 (pristine) bacteria as well as to determine if
SMX selected for a shift in a banding patterns that resembled F605 or F411.

JJM smx
Figure 5.9 RFLP comparisons of SMX treatments 0,0.004, and 0.04 pM and
untreated F411 and F605 bacterial communities using Mspl and HinPII.
There are 3 replicates (Rl, R2, and R3) for each treatment type.

fill 0 04 liM f(0S
Figure 5.10 RFLP comparisons of SMX treatments 0.04 pM R3 and untreated F411
and F605 bacterial communities
Lines on the right of the figure correspond to the high-intensity bands of 0.04 pM
SMX labled with the sample (F411 or F605) that shares a corresponding band at the
same weight (bp).
The RFLP profiles illustrate marked differences in banding patterns from the
control (0R1, R2, and R3) upon exposure to 0.04 pM SMX (0.04 R3) (Fig. 5.9, 5.10).

The exposure of 0.04 pM SMX resulted in the appearance of new bands at ~800, 400,
and 300 bp and the reduction of band intensities at ~ 150 bp compared to the
control and 0.004 pM SMX (Fig. 5.9). Additionally, the banding pattern of 0.04 R3
shares bands at 800 and 400 bp with F411 that are not present in F605 while a band
at 300 bp is shared with F605 but not present in F411 (Fig. 5.10). Unfortunately,
RFLPs of a few samples (0 R1 and 0.04 R2) did not produce high intensity banding
patterns. Although PCR product was generated prior to the RFLP, the appearances
of bands were faint. An increase in PCR product followed by normalization of DNA
prior to restriction-digest would help this. In addition, the DNA ladder (Hin IV) and
sample 0.04 R1 were distorted. A previous RFLP of the same samples (RFLP not
shown) provided an adequate result for sample 0.04 Rl, however the banding
pattern did not include resolvable bands at ~ 800 and 400 bp but was otherwise
similar to 0.04 R3. These banding patterns suggest that the environmentally
reflective concentration of 0.04 pM SMX may have the potential to alter bacterial
community composition from a pristine groundwater community. However, DNA
sequencing is needed to provide quantitative information regarding the species
composition between the three SMX treatment types (0, 0.004, and 0.04 pM) and
the SMx-impacted (F411) vs the pristine (F605) community.

6. Conclusions
The purpose of this study was to determine the effects of
environmentally reflective SMX concentrations upon the growth, denitrification,
and community composition of bacterial populations taken from a pristine
aquifer not previously exposed to SMX. The growth experiment, conducted
under denitrification conditions, illustrated that the presence of SMx(in
concentrations as low as 0.005 pM, and reflective of contaminated groundwater)
affected cell division by lengthening the overall lag phase. However, population
growth rates were not significantly decreased until the exposure of 1 pM SMX, a
concentration reflective of MIC of common pathogens.
The concurrent denitrification determinations illustrated that
denitrification rates also appeared to be impaired by the presence of 1 pM SMX.
However, this relationship seemed to be as a result of reduced growth rates
rather than a direct impact from SMX on denitrification-related enzymatic
activity. In contrast to the effect SMX had on denitrification rates, the control
consumed the highest quantity of N03 and produced the highest quantity of
N02than all SMx-treated samples over the duration of the assay. This suggests
that the control had the highest denitrification potential than all SMx-treated

samples. Additionally, as SMX concentrations increased the denitrification
potential decreased. It is unknown why the control had the highest
denitrification potential than all SMx-treated samples. It is also unknown if
denitrification potentials related to differences in growth behavior, such as
increased lag phases and single exponential phases in SMx-treated samples
compared to the two periods of growth observed in the control. Regardless, the
potential that SMx-exposure in pristine groundwater environments could impair
total IMO3' removal is an issue that should be investigated further.
When pristine groundwater bacteria were exposed to 0.04 pM SMx(a
concentration reflective of wastewater effluent and contaminated soils) without
the addition of nutrient amendments, the community structure appeared to
shift. At this concentration, the RFLP pattern differed from the pattern
represented by the control. Interestingly, the banding pattern from the 0.04 pM
SMx-treated sample differed from the control and the native pristine bacterial
community by the appearance of bands shared only with the groundwater
bacterial community that had been previously exposed to SMx and treated
wastewater. However, in order to confirm a change in community structure and

the potential relevance of that change, DNA sequencing from each community
and treatment type is required.

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It is important to address the limitations of the experimental design and
analytical methods used in this study as they impact any interpretations that can be
made. The limitations to the construction of the incubations and the analytical
methods are outlined below.
The denitrification assay, which provided information on the effects of SMx
on bacterial growth and denitrification, did not allow for the incubation of all
groundwater bacteria or even all the groundwater denitrifiers in the F605
community. This assay utilized tryptic soy broth, a general medium, with the aim
that it would stimulate the growth of a diverse abundance of bacteria, including
denitrifyers. However, denitrification has been known to occur with the use of other
non-carbon electron donors (such as Fe2+ or H2). Thus, the community of
groundwater denitrifiers might be quite complex and the use of tryptic soy broth,
albeit at O.OOlx, may have selected for only certain denitrifiers out of the whole
denitrifiying community.

The F605 groundwater assay, which provided information on the effects of
SMx on bacterial community profiles from a pristine groundwater community, also
had its limitations. In this assay, groundwater was sterilized with the use of an
autoclave and then spiked with 1.5 mL of non-sterilized groundwater and allowed to
incubate for ~ 3 weeks at room temperature. The results from this assay may not be
reflective of how SMx effects native groundwater bacteria as: 1) the incubation was
done at room temperature whereas the in-situ temperature was ~ 10 C, 2) the use
of the autoclave to sterilize the groundwater may have released carbon sources that
were previously unavailable in the groundwater, and 3) the resulting community
profiles represented the bacteria from F605 that were able to grow within 3 weeks
under these conditions, not the bacterial profile of the natural F505 community.
Analytical Methods
DAPI: DAPI is a fluorescent stain that strongly binds to DNA and has the ability to
stain metabolically active and fixed cells. Thus, DAPI does not distinguish between
live and dead cells. Although the reported cell concentrations potentially include
dead cells, overall bacterial growth implies an overall abundance of live cells.

Denitrification: Denitrification studies typically report data for all the chemicial
constituents in the denitrification process I\I03', N02', N20, and N2. However,
sometimes denitrification does not go to completion (ISI2). This will occur if the
bacterial community lacks the genes required for the completion of denitrification or
there is not enough electron donors present to allow completion. Complete
denitrification requires nitrate reductases, nitrite reductases, and nitrous oxide
reductases. The samples in this study were not analyzed for the presence of all of
these products. However, in this study it is more likely that there was not enough
electron donor (tryptic soy) present that denitrification stalled at N02".