Prepared in cooperation with the San Antonio Water System

Water-Quality Observations of the San Antonio Segment of 

the Edwards Aquifer, Texas, With an Emphasis on Processes 

Influencing Nutrient and Pesticide Geochemistry and Factors 

Affecting Aquifer Vulnerability, 2010–16

Scientific Investigations Report 2018–5060

U.S. Department of the Interior

U.S. Geological Survey
Front cover, Monitoring equipment at Frio River at Concan, Tex.

Back cover:

Top left, Pump hoist truck at Shavano Park at Fawn Drive.

Top right, Groundwater sampling equipment at AY-68-27-610 (Parkwood Park).

Middle, Comal 1 spring run.

Bottom, Nitrate sensor at the Frio River at Concan, Tex.
Water-Quality Observations of the San 

Antonio Segment of the Edwards Aquifer, 

Texas, With an Emphasis on Processes 

Influencing Nutrient and Pesticide 

Geochemistry and Factors Affecting 

Aquifer Vulnerability, 2010–16

By Stephen P. Opsahl, MaryLynn Musgrove, Barbara J. Mahler, and 

Rebecca B. Lambert

Prepared in cooperation with the San Antonio Water System

Scientific Investigations Report 2018–5060

U.S. Department of the Interior

U.S. Geological Survey
U.S. Department of the Interior

RYAN K. ZINKE, Secretary

U.S. Geological Survey

James F. Reilly II, Director 

U.S. Geological Survey, Reston, Virginia: 2018

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Suggested citation:

Opsahl, S.P., Musgrove, M., Mahler, B.J., and Lambert, R.B., 2018, Water-quality observations of the San Antonio 

segment of the Edwards aquifer, Texas, with an emphasis on processes influencing nutrient and pesticide 

geochemistry and factors affecting aquifer vulnerability, 2010–16: U.S. Geological Survey Scientific Investigations 

Report 2018–5060, 67 p., https://doi.org/10.3133/sir20185060.

ISSN 2328-031X (print)

ISSN 2328-0328 (online)
iii

Contents

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

Introduction.....................................................................................................................................................1

Purpose and Scope ..............................................................................................................................4

Hydrogeologic Framework..................................................................................................................4

Regional Land Cover.............................................................................................................................6

Methods...........................................................................................................................................................6

Study Design..........................................................................................................................................6

Rainfall Data...........................................................................................................................................7

Streamflow Measurements ................................................................................................................7

Continuous Water-Level Altitude Measurements...........................................................................7

Continuous Water-Quality Monitoring...............................................................................................7

Discrete Water-Quality Sample Collection.....................................................................................10

Analytical Methods.............................................................................................................................11

Quality Assurance for Water Samples............................................................................................13

Climate Conditions.......................................................................................................................................20

Hydrologic Conditions.................................................................................................................................22

Surface Water.....................................................................................................................................22

Groundwater........................................................................................................................................23

Geochemical Conditions.............................................................................................................................26

Meteoric Water Line ..........................................................................................................................26

Water-Rock Interaction .....................................................................................................................26

Major-Ion Distributions......................................................................................................................29

Rainfall.........................................................................................................................................29

Surface Water............................................................................................................................29

Groundwater...............................................................................................................................31

Nutrient Geochemistry................................................................................................................................34

Rainfall..................................................................................................................................................34

Surface Water.....................................................................................................................................34

Groundwater........................................................................................................................................36

Continuous Nitrate Monitoring.........................................................................................................40

Surface Water............................................................................................................................40

Unconfined Rural Groundwater ..............................................................................................41

Unconfined Urban Groundwater.............................................................................................41

Distinguishing Nitrate Sources ........................................................................................................44

Factors Affecting Nitrate ..................................................................................................................44

Nitrogen Inputs and Cycling in the Edwards Aquifer ...................................................................47

Pesticide Geochemistry..............................................................................................................................48

Rainfall..................................................................................................................................................48

Surface Water.....................................................................................................................................50

Groundwater........................................................................................................................................50

Factors Affecting Pesticides.............................................................................................................50

Pesticide Inputs and Cycling in the Edwards Aquifer ..................................................................57
iv

Factors Affecting Aquifer Vulnerability ...................................................................................................58

Land Cover...........................................................................................................................................58

Aquifer Hydrogeology ........................................................................................................................58

Changes in Hydrologic Conditions...................................................................................................58

Summary........................................................................................................................................................59

References Cited..........................................................................................................................................60

Figures

1. Map showing hydrogeologic setting and site locations in the San Antonio

segment of the Edwards aquifer, south-central Texas ..........................................................2

2. Diagram showing north-northwest-to-south-southeast section of

hydrogeologic framework and generalized groundwater flow directions,

Edwards Plateau to Gulf Coastal Plain, San Antonio region, Texas.....................................5

3. Graphs showing time series (July 2013–December 2016) of Palmer Drought

Severity Index for the Texas Edwards Plateau; spring discharge for Comal

Springs and water-level altitude for J-17 well; and daily rainfall recorded at

Rainfall data 3 .............................................................................................................................21

4. Graph showing daily median discharge, specific conductance, and

sample-collection dates at Frio River (station 0819500), San Antonio segment

of the Edwards aquifer, south-central Texas, July 2013–December 2016.........................22

5. Graphs showing water-level altitude, specific conductance, and

sample-collection times at Seco Well, Parkwood well, Shavano well,

Zarzamora well, Turtle well, Mission well, and Comal 1 spring in the San

Antonio segment of the Edwards aquifer, south-central Texas, July

2013–December 2016 .................................................................................................................24

6. Graphs showing relation between deuterium and oxygen isotopes measured

in rainfall samples, surface-water samples, and groundwater samples

collected from the San Antonio segment of the Edwards aquifer,

south-central Texas, 2013–16 ...................................................................................................27

7. Graphs showing relation between strontium-87/strontium-86 isotopic ratio

and magnesium to calcium molar ratio and time series of

strontium-87/strontium-86 isotopic ratio measured in groundwater samples

collected from the San Antonio segment of the Edwards aquifer,

south-central Texas, 2013–16....................................................................................................28

8. Graphs showing concentrations of calcium, magnesium, sodium,

bicarbonate, chloride, sulfate, and strontium measured in groundwater

samples and Comal 1 spring samples collected from the San Antonio segment

of the Edwards aquifer, south-central Texas, July 2013–December 2016.........................32

9. Graphs showing concentrations of nitrate-nitrogen and

orthophosphate-phosphorous measured in water samples collected from

groundwater wells and Comal 1 spring, San Antonio segment of the Edwards

aquifer, south-central Texas, July 2013–December 2016.....................................................39

10. Graph showing daily median discharge and daily median nitrate-nitrogen

concentration at Frio River (station 0819500), San Antonio segment of the

Edwards aquifer, south-central Texas, September 2015–December 2016........................40
v

11. Graph showing instantaneous discharge, nitrate-nitrogen concentration, and

specific conductance for a single stormwater-runoff event at Frio River

(station 0819500), San Antonio segment of the Edwards aquifer, south-central

Texas, September 20, 2016–October 20, 2016 ........................................................................41

12. Graphs showing rainfall, daily median water-level altitude, and daily median

nitrate-nitrogen for Seco well, July 2014–December 2016, and Parkwood well,

September 2014–December 2016, San Antonio segment of the Edwards

aquifer, south-central Texas .....................................................................................................42

13. Graphs showing stable nitrogen and oxygen isotopes of nitrate (delta

nitrogen-15 of nitrate and delta oxygen-18 of nitrate) measured in water

samples collected from rainfall, surface-water, and groundwater sampling

sites in the San Antonio segment of the Edwards aquifer, south-central Texas,

2011–16 .........................................................................................................................................45

14. Graph showing stable nitrogen isotopes of nitrate (delta nitrogen-15 of

nitrate) values versus nitrate-nitrogen concentration measured in

groundwater water samples collected from sampling sites in the San Antonio

segment of the Edwards aquifer, south-central Texas, 2013–16.........................................46

15. Graphs showing concentrations of ammonia-nitrogen and dissolved organic

nitrogen measured in water samples collected from sampling sites in the San

Antonio segment of the Edwards aquifer, south-central Texas, 2010–16..........................46

16. Graph showing the sum of dissolved organic nitrogen and ammonia-nitrogen

concentration versus nitrate-nitrogen concentration measured in water

samples collected from sampling sites in the San Antonio segment of the

Edwards aquifer, south-central Texas, 2010–16 ....................................................................48

17. Graph showing concentrations of the herbicide metolachlor measured in

surface-water and rainfall samples collected from sites in the San Antonio

segment of the Edwards aquifer, south-central Texas, 2012–16.........................................52

18. Graph showing a comparison of the maximum concentrations of the five

pesticides that were measured in both rainfall and surface-water samples

collected from sites in the San Antonio segment of the Edwards aquifer,

south-central Texas, 2012–16....................................................................................................52

19. Graph showing daily median water-level altitude and concentrations of the

atrazine degradate deethylatrazine (DEA) measured in water samples from

Seco well in the San Antonio segment of the Edwards aquifer, south-central

Texas, 2013–16 ............................................................................................................................53

20. Graphs showing daily median water-level altitude and concentrations of

deethylatrazine (DEA), atrazine, and simazine measured in water samples

from urban wells including Parkwood well and Zarzamora well in the San

Antonio segment of the Edwards aquifer, south-central Texas, 2013–16..........................54

21. Graph showing daily mean discharge at Comal Springs and concentrations of

deethylatrazine (DEA), atrazine, simazine, and prometon measured in water

samples collected from Comal 1 spring in the San Antonio segment of the

Edwards aquifer, south-central Texas, 2013–16 ....................................................................55

22. Graph showing daily median water-level altitude at Parkwood well and

pesticide concentrations measured in groundwater samples collected

from urban wells (Parkwood and Zarzamora wells) for atrazine and

simazine in the San Antonio segment of the Edwards aquifer, south-centra

 Texas, 2013–16............................................................................................................................56
vi

Tables

1. Sites where hydrologic data were obtained and water-quality samples were

collected in the San Antonio segment of the Edwards aquifer, south-central

Texas, 2010–16...............................................................................................................................8

2. Measured constituents, laboratory reporting levels, units, method references,

and analyzing laboratories for water-quality samples collected from sites in

the San Antonio segment of the Edwards aquifer, south-central Texas,

2010–16 .........................................................................................................................................12

3. Summary of quality-assurance data for selected compounds detected in

equipment blank and field blank water-quality samples collected from sites in

the San Antonio segment of the Edwards aquifer, south-central Texas,

2013–16 .........................................................................................................................................14

4. Summary of quality-assurance data of paired environmental and split

replicate water-quality samples and associated relative percent difference

calculations for water-quality sample pairs collected from surface-water,

spring, and groundwater sites in the San Antonio segment of the Edwards

aquifer, south-central Texas, 2013–16 .....................................................................................16

5. Summary of selected major ions measured in rainfall samples collected from

Frio River and Urban rainfall 1, 2, and 3, San Antonio segment of the Edwards

aquifer, south-central Texas, 2012–16 .....................................................................................29

6. Summary of selected major ions measured in water samples collected from

Frio River, Seco Creek, Culebra Creek, and Helotes Creek, San Antonio

segment of the Edwards aquifer, south-central Texas, 2010–16.........................................30

7. Summary of selected nutrients and stable nitrogen and oxygen isotopes of

nitrate (delta nitrogen-15 in nitrate and delta oxygen-18 in nitrate) measured

in rainfall samples collected from Frio River (representative of rural land

cover) and Urban rainfall 1, 2, and 3, San Antonio segment of the Edwards

aquifer, south-central Texas, 2012–16 .....................................................................................34

8. Summary of selected nutrients and stable nitrogen and oxygen isotopes of

nitrate (delta nitrogen-15 in nitrate and delta oxygen-18 in nitrate) measured

in surface-water samples collected from Frio River, Seco Creek, Culebra

Creek, and Helotes Creek, San Antonio segment of the Edwards aquifer,

south-central Texas, 2010–16....................................................................................................35

9. Summary of selected nutrients and stable nitrogen and oxygen isotopes of

nitrate (delta nitrogen-15 in nitrate and delta oxygen-18 in nitrate) measured

in samples collected from groundwater wells and Comal 1 spring, San

Antonio segment of the Edwards aquifer, south-central Texas, 2013–16..........................37

10. Summary of pesticide concentrations measured in rainfall samples collected

from Frio River and Urban rainfall 1, 2, and 3 in the San Antonio segment of

the Edwards aquifer, south-central Texas, 2014–16..............................................................49

11. Summary of pesticide detections in rainfall, surface-water, and groundwater

samples collected from sites in the San Antonio segment of the Edwards

aquifer, south-central Texas, 2012–16 .....................................................................................51
vii

Conversion Factors

U.S. customary units to International System of Units

Multiply By To obtain

Length

inch (in.) 2.54 centimeter (cm)

inch (in.) 25.4 millimeter (mm)

foot (ft) 0.3048 meter (m)

mile (mi) 1.609 kilometer (km)

mile, nautical (nmi) 1.852 kilometer (km)

Area

square mile (mi2

) 2.590 square kilometer (km2) 

Volume

ounce, fluid (fl. oz) 0.02957 liter (L) 

pint (pt) 0.4732 liter (L) 

quart (qt) 0.9464 liter (L) 

gallon (gal) 3.785 liter (L) 

cubic inch (in3

) 0.01639 liter (L)

cubic foot (ft3

) 28.32 cubic decimeter (dm3) 

cubic foot (ft3

) 0.02832 cubic meter (m3) 

cubic yard (yd3

) 0.7646 cubic meter (m3) 

Flow rate

cubic foot per second (ft3

/s) 0.02832 cubic meter per second (m3/s)

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

°F = (1.8 × °C) + 32

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

°C = (°F – 32) / 1.8

Datum

Vertical coordinate information is referenced to the National Geodetic Vertical Datum of 1929 

(NGVD 29) or the North American Vertical Datum of 1988 (NAVD 88).

Horizontal coordinate information is referenced to the North American Datum of 1927 (NAD 27) 

or the North American Datum of 1983 (NAD 83).

Altitude, as used in this report, refers to distance above the vertical datum.
viii

Supplemental Information

Specific conductance is given in microsiemens per centimeter at 25 degrees Celsius (μS/cm 

at 25 °C). Concentrations of chemical compounds in water are given in either milligrams per 

liter (mg/L) or micrograms per liter (μg/L). Milligrams per liter and micrograms per liter are units 

expressing the concentration of chemical compounds in solution as weight of solute (milligrams 

or micrograms) per unit volume (liter) of water.

Isotope Unit Explanations

Per mil (0/

00): A unit expressing the ratio of stable-isotope abundances of an element in a sample 

to those of a standard material. Per mil units are equivalent to parts per thousand. Stableisotope ratios are computed as follows (Kendall and McDonnell, 1998):

δX = {(Rsample – Rstandard)/Rstandard} x 1,000

where 

δ is the “delta” notation, 

X is the heavier stable isotope, and

R is ratio of the heavier, less abundant isotope to the lighter, stable isotope in a 

sample or standard. 

The δ values for stable-isotope ratios discussed in this report are referenced to the following 

standard materials.

Element R Standard identity and reference

Hydrogen Hydrogen-2/hydrogen-1 (δD) Vienna Standard Mean Ocean Water (VSMOW) 

(Fritz and Fontes, 1980)

Oxygen Oxygen-18/oxygen-16 (δ18O) Vienna Standard Mean Ocean Water (VSMOW) 

(Fritz and Fontes, 1980)

Nitrogen-nitrate Nitrogen-15/nitrogen-14 (δ15NNO3

)

USGS34 potassium nitrate (KNO3

) and 

USGS32 KNO3

 (Böhlke and others, 2003)

Oxygen-nitrate Oxygen-18/oxygen-16 of nitrate 

(δ18O-NO3

)

USGS34 KNO3

 and USGS35 sodium nitrate 

(NaNO3

) (Böhlke and others, 2003)
Water-Quality Observations of the San Antonio Segment 

of the Edwards Aquifer, Texas, With an Emphasis 

on Processes Influencing Nutrient and Pesticide 

Geochemistry and Factors Affecting Aquifer Vulnerability, 

2010–16 

By Stephen P. Opsahl, MaryLynn Musgrove, Barbara J. Mahler, and Rebecca B. Lambert

Abstract

As questions regarding the influence of increasing 

urbanization on water quality in the Edwards aquifer are 

raised, a better understanding of the sources, fate, and 

transport of compounds of concern in the aquifer—in 

particular, nutrients and pesticides—is needed to improve 

water management decision-making capabilities. The U.S. 

Geological Survey, in cooperation with the San Antonio 

Water System, performed a study from 2010 to 2016 to better 

understand how water quality changes under a range of 

hydrologic conditions and in contrasting land-cover settings 

(rural and urban) in the Edwards aquifer. The study design 

included continuous hydrologic monitoring, continuous 

water-quality monitoring, and discrete sample collection for 

a detailed characterization of water quality at a network of 

sites throughout the aquifer system. The sites were selected 

to encompass a “source-to-sink” (that is, from aquifer 

recharge to aquifer discharge) approach. Network sites were 

selected to characterize rainfall, recharging surface water, 

and groundwater; groundwater sites included wells in the 

unconfined part of the aquifer (unconfined wells) and in the 

confined part of the aquifer (confined wells) and a major 

discharging spring. Storm-related samples—including rainfall 

samples, stormwater-runoff (surface-water) samples, and 

groundwater samples—were collected to characterize the 

aquifer response to recharge.

Elevated nitrate concentrations relative to national 

background values and the widespread detection of 

pesticides indicate that the Edwards aquifer is vulnerable to 

contamination and that vulnerability is affected by factors 

such as land cover, aquifer hydrogeology, and changes in 

hydrologic conditions. Greater vulnerability of groundwater 

in urban areas relative to rural areas was evident from results 

for urban groundwater sites, which generally had higher 

nitrate concentrations, elevated δ15N-nitrate values, a greater 

diversity of pesticides, and higher pesticide concentrations. 

The continuum of water quality from unconfined rural 

groundwater sites (least affected by anthropogenic 

contamination) to unconfined urban groundwater sites (most 

affected by anthropogenic contamination) demonstrates 

enhanced vulnerability of urban versus rural land cover. 

Differences in contaminant occurrences and concentration 

among unconfined urban wells indicate that the urban parts 

of the aquifer are not uniformly vulnerable, but rather are 

affected by spatial differences in the sources of nutrients 

and pesticides. In urban areas, the shallow, unconfined 

groundwater sites showed greater temporal variability in 

both nutrient and pesticide concentrations, as well as a 

greater degree of contamination, than did deeper, confined 

groundwater sites. In comparison to that of the shallow, 

unconfined groundwater sites, the water quality of the deeper, 

confined groundwater sites was relatively invariant during 

this multiyear study. Although aquifer hydrogeology is an 

important factor related to aquifer vulnerability, land cover 

likely has a greater influence on pesticide contamination of 

groundwater. Temporal variability in hydrologic conditions 

for the Edwards aquifer is apparent in data for surface water 

as a source of groundwater recharge, water-level altitude 

in wells, spring discharge, and groundwater quality. This 

temporal variability affects recharge sources, recharge 

amounts, groundwater traveltimes, flow routing, water-rock 

interaction processes, dilution, mixing, and, in turn, water 

quality. Relations of land cover, aquifer hydrogeology, and 

changing hydrologic conditions to water quality are complex 

but provide insight into the vulnerability of Edwards aquifer 

groundwater—a vital drinking-water resource.

Introduction

The San Antonio segment of the Edwards aquifer 

(hereinafter referred to as the “Edwards aquifer”), located 

in south-central Texas (fig. 1), is a designated sole-source 

aquifer (U.S. Environmental Protection Agency, 2017a) and 

the primary water supply for more than 1.7 million people 
2 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

1,000 mg/L

Groundwater

divide

Groundwater

divide

Base modified from U.S. Geological Survey

1:2,000,000-scale and 1:24,000-scale digital data

Aquifer boundaries from 1:250,000-scale digital data

Universal Transverse Mercator, zone 14N

North American Datum of 1983

0 20 40 MILES 

0 20 40 KILOMETERS 

KERR COUNTY

REAL

COUNTY COMAL 

COUNTY

BANDERA COUNTY

BEXAR COUNTY

UVALDE COUNTY MEDINA COUNTY

EDWARDS

COUNTY KENDALL 

COUNTY

COUNTY

HAYS

KERR COUNTY

REAL

COUNTY COMAL 

COUNTY

BANDERA COUNTY

BEXAR COUNTY

UVALDE COUNTY MEDINA COUNTY

EDWARDS

COUNTY KENDALL 

COUNTY

COUNTY

HAYS

KINNEY

COUNTY

Helotes

San

Antonio

San

Antonio

BEXAR COUNTY

San Antonio

International

Airport

TEXAS

San Antonio

segment of the

Edwards aquifer

Contributing zone

Unconfined (recharge) zone

Confined zone

1

2

3

4

5

6

7

8

9

10

Map

identifier

Short name

(table 1)

Short name

(table 1)

Rainfall data 1

Rainfall data 2

Rainfall data 3

Urban rainfall 1

Urban rainfall 2

Urban rainfall 3

Frio River

Seco Creek

Culebra Creek

Helotes Creek

11

12

13

14

15

16

17

18

19

Map

identifier

Seco well

Parkwood well

Shavano well

Zarzamora well

Turtle well

Mission well

J–17 well

Comal 1 spring

Comal Springs

19

18

17

16

15

14

13

12

11

10

9

8

7

6 5

4 3

2

1

7

Site type and map identifier

 Groundwater well

 Spring—Tail points in direction of flow

 Weather station—Precipitation measurement

 Weather station—Unspecified measurement

 Streamflow-gaging station

Contributing zone to the Edwards aquifer

Edwards aquifer

 Unconfined (recharge) zone

 Confined zone

Groundwater divide

Freshwater/saline-water interface

 (1,000 milligrams per liter dissolved 

 solids concentration) (Schultz, 1994)

EXPLANATION

1

9

5

18

17

99°W

100°W

30°N

29°N

98°W

Sabinal River

Nueces River

Frio River

We st Nueces River

Guadalupe River

Medina River

Blanco River

San Antonio

River

Cibolo

Creek

Culebra

Creek

Helotes

Creek

Seco Creek

Comal

Springs

San Marcos

Springs 

Figure 1. Hydrogeologic setting and site locations in the San Antonio segment of the Edwards aquifer, south-central Texas. 

in the City of San Antonio and surrounding area (Greater 

Edwards Aquifer Alliance, 2014; Tremallo and others, 2015). 

The Edwards aquifer is both culturally and economically 

important to the region and provides recreational activities 

associated with springs and rivers (Edwards Aquifer Authority, 

2017a). The Edwards aquifer also includes habitat that 

supports endemic species such as aquifer-dwelling animals 

and spring-dependent plants, many of which are classified 

as threatened or endangered (Edwards Aquifer Authority, 

2017a). Water quality, recreational activities, and habitat of 

the Edwards aquifer can be compromised by the presence of 

various compounds of concern; therefore, there is a need to 

identify compounds of concern and better understand their 

sources, fate, and transport throughout the aquifer.

The Edwards aquifer is typical of karst aquifers, which 

are composed of soluble host rocks, usually carbonate rocks 

that have dissolved preferentially to form large voids and 

conduits (White, 1988). Karst features are characteristic of 

the Edwards aquifer and make the system susceptible to 

contamination because of the rapid transport of groundwater 

through high-porosity voids and conduits with little 

opportunity for subsurface filtration, sorption, or degradation 

of dissolved or particulate compounds (White, 1988). The 

Edwards aquifer recharge zone is particularly susceptible 

to contamination from the land surface because of the 

dynamic and rapid recharge that occurs through karst features 

such as sinkholes and losing streams during groundwater 

recharge events (Sharp and Banner, 1997). Dynamic recharge 

affects the timing and magnitude of changes in water 

quality—in particular nutrients and pesticides derived from 

anthropogenic sources. 
Introduction 3

Previous water-quality studies of the Edwards aquifer 

have documented spatial and temporal variability in nutrient 

concentrations across the aquifer (Musgrove and others, 

2010, 2011; Opsahl and others, 2017), which raises questions 

regarding the origin of nutrients. Nutrients, essential 

elements for plant and animal life, can have both natural 

and anthropogenic sources including rainfall, decomposing 

plants, soil erosion, human and animal wastes, inorganic 

and organic fertilizer, treated wastewater, and raw sewage 

(Dubrovsky and others, 2010). Nutrients can have deleterious 

effects on water quality especially if present at concentrations 

substantially elevated above natural levels. Nitrate (NO3

) in 

drinking water is a regulated contaminant with a maximum 

contaminant level of 10 milligrams per liter (mg/L) (U.S. 

Environmental Protection Agency, 2017b). Excess nitrogen 

(N) in the form ammonia (NH4

) and NO3 in surface water 

promotes algal blooms (eutrophication), and decomposition of 

the algae consumes dissolved oxygen, causing other aquatic 

organisms to die (Rabalais and others, 2002; Ansari and 

others, 2010). The types and amounts of nutrients that might 

be introduced into the aquifer during groundwater recharge 

will likely vary because of land-use practices, among other 

factors. For example, nutrient pollution in the form of NO3

might be attributed to fertilizer in both rural (agriculture) and 

urban (landscaping) settings (Dubrovsky and others, 2010). In 

the Edwards aquifer, the proportion of nutrients derived from 

rural and urban sources relative to contributions from natural 

sources remains uncertain.

NO3

 isotope values (delta nitrogen-15 of nitrate 

[δ15N-NO3

] and delta oxygen-18 of nitrate [δ18O-NO3]) have 

been used to identify sources of NO3

, distinguish groundwater 

flow paths, and provide information on biogeochemical 

processes that are relevant to tracing NO3

 sources and sinks 

in groundwater systems (Clark and Fritz, 1997; Kendall, 

1998; Kendall and others, 2014). The systematic relation 

between δ15N-NO3

 and δ18O-NO3 in common NO3 sources, 

compiled from numerous studies, provides insight into sources 

of NO3

 contributing to environmental samples (Kendall and 

others, 2014). For example, NO3

 in synthetic fertilizers is 

distinguished from human and animal waste on the basis of 

lower δ15N-NO3

 values and higher δ18O-NO3 values. NO3

in rainfall is clearly distinguished from other NO3

 sources 

because of its high δ18O-NO3

 values and low δ15N-NO3 values 

relative to surface-water and groundwater values. δ15N-NO3

values that range from about 3 to 9 per mil are generally 

considered representative of a soil NO3

 source. Numerous 

studies, however, have indicated that δ15N-NO3

 values in 

environmental samples between 3 and 10 per mil might result 

from a mixed source of possible contributors that include 

inorganic fertilizers, human and animal waste, and natural 

soil NO3

 (Heaton, 1986; Wells and Krothe, 1989; Katz and 

others, 2004; Oren and others, 2004; Choi and others, 2007; 

Kendall and others, 2014). A broader understanding of the 

distribution of all N species and the associated biogeochemical 

transformations related to the formation and breakdown 

of NO3

 can potentially enhance the use of δ15N-NO3 and 

δ18O-NO3

 values for further determining potential NO3 sources 

in the Edwards aquifer. 

The term “pesticides,” as used in this study, refers to 

anthropogenic organic compounds and their degradates 

designed to kill or control plants (herbicides), insects 

(insecticides), fungi (fungicides), or other organisms, which 

are widely used in both rural and urban settings. Since the 

1970s, less-persistent pesticides have been developed, and 

although they degrade more quickly than their predecessors, 

they are more soluble and frequently are detected in 

groundwater (Gilliom and others, 2006). Agricultural use of 

pesticides exceeds nonagricultural use, but nonagricultural 

use has been increasing since 1998 (Gilliom and others, 

2006). Although pesticides are used in both rural and urban 

settings, the distribution of specific pesticides typically 

differs between rural and urban sources (Van Metre and 

others, 2016). Between 1992 and 2001, the herbicides most 

commonly detected in urban areas in the United States were 

prometon (streams) and atrazine (groundwater), and the 

insecticides most commonly detected in urban areas were 

diazinon (streams) and dieldrin (groundwater) (Gilliom 

and others, 2006). In agricultural areas, the herbicides most 

commonly detected between 1992 and 2001 were atrazine 

(streams) and its degradate deethylatrazine (also known as 

CIAT [2-Chloro-4-isopropylamino-6-amino-s-triazine] and 

hereinafter referred to as “DEA”) (groundwater), and the 

insecticides most commonly detected were diazinon (streams) 

and carbofuran (groundwater) (Gilliom and others, 2006). 

However, the intensity of use of individual chemicals changes 

as new pesticides are introduced, as land-use practices change, 

and when their use is banned or curtailed. Data on pesticide 

use in agricultural areas (Gilliom and others, 2006) indicate 

that since 2000 only the use of atrazine has remained relatively 

steady—by 2014 use of prometon, diazinon, and carbofuran 

had decreased. Conversely, use of other herbicides, such 

as glyphosate, and insecticides, such as imidacloprid, has 

increased since 2000 (Gilliom and others, 2006). 

Previous studies have documented the occurrence of 

pesticides in the Edwards aquifer (Bush and others, 2000; 

Fahlquist and Ardis, 2004; Musgrove and others, 2010). The 

pesticides most frequently detected throughout the aquifer 

in those studies were simazine, prometon, atrazine, and 

the atrazine degradate DEA, and these were detected most 

frequently in samples of unconfined groundwater collected at 

shallow urban wells. Musgrove and others (2010) reported that 

temporal variability in concentrations of these compounds was 

inconsistent among five wells with eight samples collected 

over 10 years, but that elevated concentrations of atrazine 

generally coincided with elevated water-level altitudes (waterlevel altitude is hereinafter referred to as “WLA”).

Although the occurrence of nutrients and pesticides 

in both the unconfined and confined groundwater of the 

Edwards aquifer, as well as in surface-water recharge, has 

been documented (Bush and others, 2000; Fahlquist and 

Ardis, 2004; Musgrove and others, 2010; Opsahl, 2012), 

prior studies have not been designed to distinguish between 
4 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

different sources (rural versus urban), nor have they addressed 

how changes in hydrologic conditions influence water quality 

in the aquifer. Identifying geographic sources of nutrients and 

pesticides to the aquifer is further complicated by atmospheric 

inputs in rainfall, which can contribute both nutrients and 

pesticides with uncertain origins that may be either close to 

or distant from areas receiving rainfall (Nations and Hallberg, 

1992; Hüskes and Levsen, 1997; Miller and others, 2000; 

Lajtha and Jones, 2013). The connectivity between water 

quality in groundwater recharge and deeper, confined water 

within the aquifer also has not been well characterized. 

In particular, understanding the role of surface-water and 

groundwater interaction and its effects on water quality under 

a range of hydrologic conditions, especially during periods 

of rapid groundwater recharge, is needed to inform that 

understanding. To address these questions, the U.S. Geological 

Survey, in cooperation with the San Antonio Water System, 

performed a study from 2010 to 2016 to better understand 

how water quality changes under a range of hydrologic 

conditions and in contrasting land-cover settings (rural and 

urban) in the Edwards aquifer. The wide range of hydrologic 

conditions that occurred during the study, from drought to 

wetter than normal, and geochemical processes occurring 

in the aquifer affected the occurrence, fate, and transport of 

nutrients and pesticides. 

Purpose and Scope

The purpose of this report is to assess changes in selected 

water-quality constituents across the Edwards aquifer with an 

emphasis on the processes influencing nutrient and pesticide 

geochemistry and the factors affecting aquifer vulnerability. 

Differences in nutrients and pesticide concentrations during 

dry and wet conditions and for periods of base flow and 

stormwater runoff are described for rainfall, surface-water, 

and groundwater samples collected during 2010–16. The 

effects of different types of land cover (rural and urban) on 

the differences between water-quality constituents in surfacewater and groundwater samples are explored. Rural and urban 

sources of pesticides are evaluated along with how changes in 

hydrologic conditions influence water quality in the aquifer. 

Changes in WLAs between confined and unconfined parts of 

the aquifer are described in the context of changes in water 

quality and aquifer vulnerability.

Hydrogeologic Framework

The Edwards aquifer developed in heavily faulted, 

fractured, and karstified early Cretaceous-aged limestone 

and dolomitic rocks with extensive dissolution that yielded 

a porous matrix (Barker and Ardis, 1996). The aquifer is 

characterized by relatively high transmissivities (Lindgren, 

2006). Typical of karst aquifers, most groundwater storage 

occurs within the aquifer matrix, but most transport occurs 

within conduits, which often dominate groundwater flow 

where present (White, 2002). The aquifer is present in a 

narrow band in south-central Texas—bounded at its western 

edge by a groundwater divide in Kinney County and at the 

northeastern edge by a groundwater divide in Hays County 

(fig. 1). Late Cenozoic faulting along the Balcones fault zone 

(fig. 2) formed a series of high-angle normal en echelon faults 

(Maclay and Small, 1983). The faulting resulted in a series of 

blocks of Edwards aquifer rocks that are offset and divide the 

confined and unconfined parts of the aquifer. The recharge 

zone (unconfined part of the aquifer) consists of about 

1,250 square miles (mi2) of Edwards Group limestone exposed 

at the surface (fig. 2) (Hamilton and others, 2008). The 

northern limit of the recharge zone (outcrop) defines the 

northern aquifer boundary (fig. 1). Downdip of the recharge 

zone, the aquifer becomes confined and is artesian. The 

aquifer is bounded to the south by the freshwater/saline-water 

interface, downgradient of which groundwater salinity rises 

rapidly (dissolved solids concentration exceeding 1,000 mg/L), 

and the water is considered unpotable (Schultz, 1994). 

The majority of recharge to the Edwards aquifer occurs 

by direct infiltration from streams that drain the contributing 

zone (fig. 1) as they cross the recharge zone and discharge 

into the aquifer through open-solution channels (Puente, 1978; 

Maclay and Land, 1988). Intense storms with heavy rainfall 

are common, and flash flooding in streams greatly enhances 

groundwater recharge following storm events. Estimates of 

recharge from losing streams range from 60 to 80 percent 

(Klemt and others, 1979; Maclay and Land, 1988; Thorkildsen 

and McElhaney, 1992; Ockerman, 2005). Recharge is not 

evenly distributed across the aquifer; most recharge occurs 

from streams located in the western part of the aquifer 

(Puente, 1978). The remaining aquifer recharge comes from 

a combination of direct infiltration through the recharge zone 

and leakage from the underlying Trinity aquifer (fig. 2) into 

the Edwards aquifer (Sharp and Banner, 1997; Lindgren and 

others, 2004). Groundwater pumping in the San Antonio area 

has increased about fivefold since the 1930s, but WLA and 

spring discharge have not shown long-term declines because 

the aquifer readily recharges during periods of rainfall (Bush 

and others, 2000; Edwards Aquifer Authority, 2015).

The flow paths of recharge and regional groundwater 

in the Edwards aquifer are described in Maclay and Land 

(1988) and Clark and Journey (2006). Recharge water in 

the unconfined zone flows downdip to the confined zone 

(fig. 1). Regional groundwater flow is generally west to east 

from Kinney County through Uvalde, Medina, and Bexar 

Counties. The regional flow path shifts northeast through 

Comal and Hays Counties, where natural discharge occurs 

at large springs, predominantly Comal Springs and San 

Marcos Springs (Lindgren, 2006; Musgrove and Crow, 2012). 

Pumping for agriculture and municipal purposes also is a 

major component of discharge from the Edwards aquifer 

(Hamilton and others, 2008). 
Balcones

escarpment

NOT TO SCALE

Trinity Group

Edwards aquifer

recharge zone

EXPLANATION

Upper confining unitEdwards aquifer Unconfined aquifer Confined aquiferEdwards-Trinity aquiferTrinity aquifer

LimestoneSandUnderlying rock unitsApproximate freshwater/saline- water interfaceStreamGeneralized conduit systemGeneralized direction of groundwater flowFault—Arrows indicate direction of movement

NORTH-NORTHWEST 

Edwards

Trinity Group

Edwards Group

Trinity

Group

Del Rio Clay, 

Eagle Ford Group, Austin C

and overlyi

Trinity Group

Edwards Group

Hammett Shale

Group

Catchment area

Unconfined

aquifer Confined aquifer

EDWARDS PLATEAU

HILL COUNTRY

GU

Balcones fault zone

Streamflow

diminishes

to zero 

L a n d s u r f a c e

Figure 2. North-northwest-to-south-southeast section of hydrogeologic framework and generalized groundwater flow directions, Edwards Plateau t

Antonio region, Texas (from Musgrove and others, 2011, modified from Barker and Ardis, 1996, plates 1 and 3).
6 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Regional Land Cover

As discussed in Musgrove and others (2010), regional 

land cover in the study area generally correlates with 

physiography. The contributing zone of the Edwards Plateau 

(fig. 1; shown as the catchment area in fig. 2) is characterized 

by thin soils and is predominantly undeveloped rangeland. 

There is little agriculture; agricultural lands are mostly to 

the west of San Antonio, over the confined aquifer, and 

what little agricultural land there is has decreased in recent 

years (Musgrove and others, 2016). Regional land cover 

is characterized as 3 percent agriculture, 6 percent urban, 

and 90 percent forest and rangeland (including shrub and 

grassland); the remaining 1 percent is water, wetlands, 

and barren land (Homer and others, 2001). San Antonio 

is the seventh largest city in the United States with a 2016 

population of about 1.5 million (U.S. Census Bureau, 2017). 

The City of San Antonio is the principal urban area and 

includes much of Bexar County in the east-central part of the 

aquifer region. Bexar County is rapidly urbanizing, especially 

on San Antonio’s north side in the aquifer’s contributing and 

recharge zones. For example, while the population of Bexar 

County between 2000 and 2010 increased by 23 percent, 

growth in several north-side communities exceeded 50 percent 

(U.S. Census Bureau, 2017). 

Methods

A monitoring network was designed to provide long-term 

geochemical data of water quality from rainfall, stream, and 

groundwater sites (fig. 1; table 1). Rainfall, surface-water, 

and groundwater samples were collected over a period of 

approximately three and a half years (July 2013 through 

December 2016) at different sites throughout the study area. 

Site selection took into account different land-cover types 

(rural and urban environments) and included unconfined 

zone (relatively shallow) and confined zone (relatively deep) 

groundwater wells (fig. 1). All sites will hereinafter be referred 

to by their short names (table 1).

Study Design

The study design built upon knowledge gained from 

previous studies of water quality in the region (including 

Mahler and others, 2006, 2011a, b; Musgrove and others, 

2010, 2011; Opsahl, 2012). The study design included 

continuous hydrologic and water-quality monitoring (stream 

discharge, WLA, specific conductance in microsiemens 

per centimeter at 25 degrees Celsius [hereinafter referred 

to as “SC”], and NO3

 concentration) and discrete sample 

collection for a detailed characterization of water quality at 

a network of sites throughout the aquifer system. The sites 

were selected to encompass a “source-to-sink” (that is, from 

aquifer recharge to aquifer discharge) approach. Network 

sites were selected to characterize rainfall, recharging surface 

water, and groundwater; groundwater sites included wells 

in the unconfined part of the aquifer (unconfined wells) 

and in the confined part of the aquifer (confined wells) and 

a major discharging spring. Stormwater-runoff (surfacewater) and storm-response (groundwater) samples also were 

included to characterize water quality in response to rainfall 

and recharge-generating storms, and thus, the frequency of 

water-quality sample collection was higher during periods of 

rapid hydrologic change. Rainfall samples were collected for 

water-quality analyses at the Frio River (representative of rural 

land cover) and at three separate sites located within the City 

of San Antonio (representative of urban land cover) (fig. 1). 

All surface-water sites are ephemeral, although the Frio River 

ceases to flow only during periods of prolonged drought. Seco 

Creek, Culebra Creek, and Helotes Creek are typically dry and 

flow only for intermittent periods following relatively large 

storm events. Rural surface-water sites included the Frio River, 

Seco Creek, and Culebra Creek. Helotes Creek was selected 

as representative of urban surface water on the basis of 

increasing urbanization in its watershed (Opsahl, 2012). Seco 

well was selected to represent unconfined groundwater in the 

rural portion of the system, and Parkwood well was selected to 

represent urban unconfined groundwater. Zarzamora well was 

selected to represent urban confined (deeper) groundwater. 

Comal 1 spring, located near the downgradient end of the 

regional flow path (fig. 1), was selected to represent an 

integrator of regional aquifer groundwater quality (Musgrove 

and Crow, 2012) and, for the purposes of this report, will 

hereinafter be considered as groundwater. In addition, selected 

sites from a complementary Enhanced Trends Network (ETN) 

study by the U.S. Geological Survey National Water Quality 

Assessment (NAWQA) project also were used in this study. 

The ETN sites, all within the City of San Antonio, include 

an unconfined urban well (Shavano well), a confined urban 

production well (Turtle well), and a deeper, confined urban 

production well that is artesian (the Mission well) located 

further downdip and close to the freshwater/saline-water 

interface (fig. 1). The placement of continuous-monitoring 

sensors considered the water chemistry of the Edwards 

aquifer, which is generally well mixed vertically, consistent 

with the porous karst properties and high transmissivities 

of the aquifer (Lindgren, 2006; Musgrove and others, 2010, 

2011). Continuous-monitoring sensors were positioned within 

the screened interval in each well (with the exception of the 

Mission well and the Turtle well), and data from these wells 

are assumed to be representative of the entire saturated zone 

for each site. Groundwater samples were collected from 

unconfined wells completed in the recharge zone of the 

aquifer and from confined wells completed in the deeper part 

of the aquifer. Groundwater wells in the unconfined zone of 

the aquifer (hereinafter referred to as “unconfined wells”) 

were selected to characterize the rapid changes in WLA and 

water quality that occur during groundwater recharge, and 

groundwater wells in the confined zone (hereinafter referred 

to as “confined wells”) were selected to represent the part of 
Methods 7

the aquifer system that responds more slowly to groundwater 

recharge. The spatial and temporal differences in sample 

collection provided an opportunity to investigate processes 

affecting the fate and transport of nutrients and pesticides 

within the Edwards aquifer. The types of data collected for 

this study included continuous monitoring of physiochemical 

parameters, continuous monitoring of water-quality 

parameters, water-quality data from discretely collected 

samples, and ancillary climatological and hydrologic data to 

help interpret the geochemical data, all of which are described 

below and summarized in table 1. Not all data collected during 

the study are included in this report; however, all data are 

available for download in a companion data release (Opsahl 

and others, 2018).

Rainfall Data

Daily rainfall data were obtained for the study period 

from meteorological sites near some of the sampled wells 

(Seco well and Parkwood well) and the San Antonio 

International Airport (fig. 1; table 1). Daily rainfall data were 

acquired from the Global Historical Climatology Network 

(GHCN) (Menne and others, 2012; National Oceanic and 

Atmospheric Administration, 2017a). During the study 

period, no rainfall data were reported for 12 days of the 

record for Rainfall data 2 (located near Seco well). GHCN 

site US1TXMDN041 (National Oceanic and Atmospheric 

Administration, 2017a), the next nearest GHCN site to the 

Rainfall data 2 and Seco well, is located about 5.6 miles (mi) 

southeast of Seco well. No rainfall was reported at GHCN 

site US1TXMDN041 for those corresponding dates with 

two exceptions where no data were available from Rainfall 

data 2 or GHCN site US1TXMDN041 on June 7, 2015, and 

December 30, 2016 (National Oceanic and Atmospheric 

Administration, 2017a). No rainfall was recorded regionally 

on June 7, 2015, or December 30, 2016 (National Oceanic 

and Atmospheric Administration, 2017b), so rainfall was 

assumed to be zero. No rainfall data were available for 16 

days of the record for Rainfall data 1 (located near Parkwood 

well). GHCN site US1TXBXR287 (National Oceanic and 

Atmospheric Administration, 2017a), the next nearest GHCN 

site to Rainfall data 1 and Parkwood well, is located about 

3.7 mi southwest of Parkwood well. Rainfall was reported for 

five of those dates, and daily rainfall values from GHCN site 

US1TXBXR287 were substituted to supplement the record 

for Rainfall data 1 (0.07 inches [in.] on January 1, 2015; 

0.02 in. on June 14, 2015; 2.30 in. on June 15, 2015; 0.25 in. 

on June 17, 2015; and 0.03 in. on January 1, 2016). No rainfall 

data were available from Rainfall data 1 or US1TXBXR287 

for June 28, 2016 (National Oceanic and Atmospheric 

Administration, 2017a). No rainfall was reported regionally 

on June 28, 2016 (National Oceanic and Atmospheric 

Administration, 2017b), so rainfall was assumed to be zero 

on June 28, 2016. The rainfall record for the San Antonio 

International Airport was complete for the study period.

Streamflow Measurements

Streamflow-gaging stations at Culebra Creek, Helotes 

Creek, and Frio River (fig. 1; table 1) are operated as part 

of an existing network of cooperatively funded USGS 

streamflow-gaging stations. Stage at each surface-water site 

was recorded every 15 minutes and transmitted hourly by the 

Geostationary Operational Environmental Satellite (GOES) 

transmitter to the USGS National Water Information System 

(NWIS) database (U.S. Geological Survey, 2017). Discharge 

for Helotes Creek and the Frio River is computed from the 

15-minute stage data by using a stage-discharge relation 

unique to each surface-water site. The streamflow-gaging 

stations are visited periodically to verify the recorded stage 

and obtain measurements of flow, used to verify the stagedischarge relation (Rantz and others, 1982a; Turnipseed and 

Sauer, 2010). The information obtained during the periodic 

site visits is applied to update the record (and stage-discharge 

relation, as needed) and to compute continuous discharge 

(Rantz and others, 1982b; Kennedy, 1983, 1984; Sauer and 

Turnipseed, 2010), which is stored in the NWIS database 

(U.S. Geological Survey, 2017). Seco Creek (fig. 1; table 1) 

is a previously established USGS streamflow-gaging station 

that was used only for periodic water-quality sample collection 

in this study. The Culebra Creek site reports only stage data 

because there are too few discharge measurements to date 

(December 2017) to develop a stage-discharge relation.

Continuous Water-Level Altitude Measurements

Continuous WLA data were collected from Seco well, 

Parkwood well, Shavano well, and Zarzamora well (fig. 1; 

table 1) by using methods described by Cunningham and 

Schalk (2011). The wells were equipped with a submersible 

pressure transducer, data recorder, and GOES transmitter. 

Water-level data were recorded in feet below a known land 

surface datum at 15-minute intervals and periodically verified 

against a calibrated water-level tape and corrected as required. 

The 15-minute water-level data were transmitted hourly via 

GOES to the NWIS database (U.S. Geological Survey, 2017). 

In NWIS, the water-level data were then converted to WLA 

data by using either the National Geodetic Vertical Datum of 

1929 (NGVD 1929) or the North American Vertical Datum of 

1988 (NAVD 1988) reference datum.

Continuous Water-Quality Monitoring

Continuous water-quality monitoring began in July 2013 

and continued through December 2016, although the exact 

beginning dates for different data types varied among sites. 

For all wells equipped with in-situ water-quality sensors 

(Seco, Parkwood, Shavano, and Zarzamora wells), placement 

was at approximate depths shown in table 1. Prior information 

for Seco well including vertical profiling of SC, temperature, 

and vertical flow logging (Opsahl and others, 2017) was used 
8 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Table 1. Sites where hydrologic data were obtained and water-quality samples were collected in the San Antonio segment of the 

Edwards aquifer, south-central Texas, 2010–16.

[GHCN, Global Historical Climatology Network; USGS, U.S. Geological Survey; EAA, Edwards Aquifer Authority; dd, decimal degrees referenced to North 

American Datum of 1983; ft, foot; NAVD 88, North American Vertical Datum of 1988; SSE, south southeast; TX, Texas; US, United States; NA, not applicable; 

RF, daily rainfall total; WSW, west southwest; QWrf, rainfall water quality; SF, streamflow; QWp, periodic water quality; QWs, stormflow-runoff water quality; 

QWc, continuous water quality; QWn, continuous nitrate water quality; FM, farm to market road; WLA, water-level altitude; --, not available]

Map 

identifier 

(fig. 1)

Site type GHCN, USGS, or EAA 

station number

GHCN, USGS, or 

EAA station name

Short 

name 

(fig. 1)

1 Weather station 1

US1TXBXR095 Helotes 3.7 SSE TX US Rainfall data 1

2 Weather station 1

US1TXMDN021 D HANIS 3.5 WSW TX US Rainfall data 2

3 Weather station 1

USW00012921 San Antonio International Airport TX US Rainfall data 3

4 Rainfall-collection station 2

293537098261500 EARZ Urban Precipitation Site 1 at San Antonio, Tex. Urban rainfall 1

5 Rainfall-collection station 2

293044098325100 EARZ Urban Precipitation Site 2 at San Antonio, Tex. Urban rainfall 2

6 Rainfall-collection station 2

88480701 USGS South Texas Program Office at San Antonio, Tex. Urban rainfall 3

7 Streamflow-gaging and 

rainfall-collection station

2

08195000 Frio River at Concan, Tex. Frio River

8 Streamflow-gaging station 2

08202600 Seco Creek at FM 1796 near D’Hanis, Tex. Seco Creek

9 Streamflow-gaging station 2

08180941 Culebra Creek in Government Canyon State Natural 

Area near Helotes, Tex.

Culebra Creek

10 Streamflow-gaging station 2

08181400 Helotes Creek at Helotes, Tex. Helotes Creek

11 Groundwater well3 2

292618099165901 Seco Creek Well Seco well

12 Groundwater well3 2

293252098380801 AY-68-27-610 (Parkwood Park) Parkwood well

13 Groundwater well3 2

293516098325501 Shavano Park at Fawn Drive Shavano well

14 Groundwater well4 2

292943098354404 AY-68-36-132 (Z DED) Zarzamora well

15 Groundwater well4 2

293116098334101 Turtle Creek 3 Turtle well

16 Groundwater well4 2

292331098294501 Mission 7 Mission well

17 Groundwater well4 5

6837203 Bexar County index well (J-17) J-17 well

18 Spring 2

294300098080001 Comal 1 Comal 1 spring

19 Spring 2

08168710 Comal Springs at New Braunfels, TX Comal Springs
Methods 9

Table 1. Sites where hydrologic data were obtained and water-quality samples were collected in the San Antonio segment of the 

Edwards aquifer, south-central Texas, 2010–16.—Continued

[GHCN, Global Historical Climatology Network; USGS, U.S. Geological Survey; EAA, Edwards Aquifer Authority; dd, decimal degrees referenced to North 

American Datum of 1983; ft, foot; NAVD 88, North American Vertical Datum of 1988; SSE, south southeast; TX, Texas; US, United States; NA, not applicable; 

RF, daily rainfall total; WSW, west southwest; QWrf, rainfall water quality; SF, streamflow; QWp, periodic water quality; QWs, stormflow-runoff water quality; 

QWc, continuous water quality; QWn, continuous nitrate water quality; FM, farm to market road; WLA, water-level altitude; --, not available]

Map 

identifier 

(fig. 1)

Latitude

(dd)

Longitude 

(dd)

Elevation of 

land surface 

(ft above 

NAVD 88)

Well depth 

(ft below land 

surface)

Waterquality 

sensor

depth 

(ft below 

land surface)

Nitrate 

sensor 

depth 

(ft below 

land surface)

Data type

1 29.51694 98.67500 NA NA NA NA RF

2 29.32611 99.33778 NA NA NA NA RF

3 29.54444 98.48389 NA NA NA NA RF

4 29.59362 98.43750 NA NA NA NA QWrf

5 29.51218 98.54738 NA NA NA NA QWrf

6 29.52967 98.49502 NA NA NA NA QWrf

7 29.48856 99.70478 NA NA NA NA SF, QWp, QWs, QWc, QWn, 

QWrf

8 29.43967 99.28489 NA NA NA NA QWp

9 29.53917 98.75139 NA NA NA NA QWs

10 29.57856 98.69169 NA NA NA NA SF, QWs

11 29.43857 99.28337 1,008.7 538 370 380 WLA, QWp, QWc, QWn

12 29.54778 98.63556 896 229 205 210 WLA, QWp, QWc, QWn

13 29.58778 98.54861 975 300 270 295 WLA, QWp, QWc, QWn

14 29.49528 98.59556 848 596.5 590 NA WLA, QWp

15 -- -- 946 550 NA NA QWp, QWc

16 -- -- 585 1,500 NA NA QWp, QWc

17 29.47917 98.43250 6730.8 874 NA NA WLA

18 29.71278 98.13750 NA NA NA NA QWp, QWc

19 29.70606 98.12251 NA NA 1 NA SF

1

GHCN station.

2

USGS station.

3

Groundwater well completed in the unconfined part of the aquifer.

4

Groundwater well completed in the confined part of the aquifer.

5

EAA station.

6

Bexar County index well (J-17) water-level altitudes are reported and referenced to the National Geodetic Vertical Datum of 1929.
10 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

to identify the most productive zone (about 360 feet [ft] below 

land surface). Vertical flow in this well was downward, and 

the sensors were placed 360 ft below land surface to monitor 

this more productive water-bearing zone. At Parkwood and 

Shavano wells, sensors were placed near the bottoms of 

these wells, which were likely to be the more water-bearing 

zones based on stratigraphy and well logs. At Zarzamora 

well, sensors were placed midway in the screened interval 

(between 586 and 597 ft below land surface), a depth typical 

of surrounding production wells. Notably, all wells are situated 

in a porous karst setting, and although detailed vertical 

profiling was not included in the study design, prior studies 

have indicated that wells in this setting are vertically well 

mixed (Lindgren, 2006; Musgrove and others, 2011; Opsahl 

and others, 2017). Shavano well, Mission well, Turtle well, 

and Comal 1 spring were instrumented with multiparameter 

EXO 1 water-quality sensors (YSI Incorporated) to collect 

water temperature, SC, dissolved oxygen concentration, and 

pH data. The Mission and Turtle wells were instrumented ex 

situ; that is, the water-quality sensors measure a stream of 

water that is diverted at the well head and passes through a 

flowthrough chamber when the wells are pumping or flowing. 

At Comal 1 spring, the water-quality sensor was placed into 

the spring orifice. Seco, Parkwood, and Zarzamora wells were 

instrumented in situ with YSI OMS600 water-quality sensors 

placed downhole in the water column to measure SC and 

water temperature. The water-quality sensors were maintained 

with periodic field visits to clean and calibrate the instruments 

by using methods described by Wagner and others (2006).

Continuous monitoring of NO3

 was implemented at Seco 

well in July 2014, Parkwood well in September 2014, and the 

Frio River in September 2015 by using submersible ultraviolet 

nitrate analyzer (SUNA) sensors (Sea-bird Scientific, 2018). 

The measuring principle of the NO3

 sensors used in this study 

is based on optical absorption (Pellerin and others, 2013). NO3

and nitrite (NO2

) have similar absorption properties, and the 

sensors detect both compounds as part of the measurement. 

However, NO2

 concentrations in the types of water monitored 

in this study are uniformly low (Opsahl and others, 2017), and 

the NO3

-sensor values reported in this study will hereinafter 

be referred to as “NO3

 concentrations.” The NO3 sensors were 

maintained with periodic field visits to clean and verify the 

calibration by using methods described by Pellerin and others 

(2013). The precision of the NO3

 sensor is plus or minus 

0.07 mg/L with a 95 percent confidence interval (Opsahl 

and others, 2017). High turbidity prevents adequate light 

transmission through the flow cell of the instrument (Pellerin 

and others, 2013), and NO3

 values recorded following storms 

at the Frio River were censored from the dataset until a 

minimum criterion for light transmission was again met.

Continuous water-quality and streamflow data were 

collected at 15-minute intervals, with the exception of NO3

, 

for which data were collected at hourly intervals; these data 

are referred to as unit values and were transmitted hourly via 

GOES to the NWIS database (U.S. Geological Survey, 2017). 

The data records for Turtle and Mission production wells 

were intermittent throughout the monitoring period, providing 

reliable data only when the wells were being pumped (Turtle) 

or flowing unimpeded (Mission). Daily median values 

for streamflow, WLA, SC, and NO3 and streamflow were 

computed from the available unit values.

Discrete Water-Quality Sample Collection

The exact period of water-quality sample collection 

varied among the different sites, but the majority of sampling 

occurred between July 2013 and December 2016. There were 

a few samples that were collected earlier than July 2013, 

including one stormwater-runoff sample from Culebra Creek 

in April 2010, four stormwater-runoff samples from Helotes 

Creek between October 2011 and May 2013, and two urban 

rainfall samples collected in January 2012 from site Urban 

rainfall 3 (table 1). Samples collected earlier than July 2013 

represent sample types that are particularly difficult to collect 

(for example, ephemeral streams, which rarely flow). These 

additional measurements were based on the same collection 

techniques and analytical methods as the rest of the samples 

in this study and are included to expand the available data for 

these sites.

During July 2013 through December 2016, the frequency 

of routine sampling for the Frio River and three groundwater 

sites (Seco, Parkwood, and Zarzamora) was about every 

6–8 weeks. Sites that were part of the ETN network (Shavano, 

Turtle, and Mission wells) were sampled from November 

2013 through December 2016. The sampling frequency at 

the ETN sites was about every 6–10 weeks, similar to the 

other sites in the network, though sometimes longer. Routine 

samples from all sites were collected without regard to 

hydrologic conditions. Storm-related samples—including 

rainfall samples, stormwater-runoff (surface-water) samples, 

and groundwater samples—were collected to characterize the 

aquifer response to recharge. These samples were collected 

opportunistically as conditions permitted. Rainfall samples 

were dependent on storm events, surface-water samples were 

dependent on increased streamflow following storm events, 

and groundwater samples were dependent on rising water 

levels following rainfall and recharge events. As a result, 

storm-related samples were collected at different times for 

different sites. 

Rainfall samples were collected by following standard 

USGS methods documented in the “National Field Manual 

for the Collection of Water-Quality Data” (U.S. Geological 

Survey, variously dated). Rainfall samples were collected at 

four sites, three in the urban San Antonio area (multiple sites 

in proximity were used to accommodate sampling logistics 

and site availability) and one in the rural western part of the 

aquifer region (fig. 1; table 1). Rainfall was collected by 

using a bulk rainfall collector consisting of a stainless steel 

funnel that emptied into a polyethylene container. The rainfall 

collectors were placed at the respective locations prior to an 
Methods 11

imminent rainfall event, and the samples were retrieved as 

soon as practical following the event. Each rainfall sample 

was a single-composite sample representing the average water 

quality of collected rainfall for the event. Samples were placed 

on ice and transported to the USGS South Texas Program 

Office Laboratory in San Antonio for processing.

Routine and storm-runoff samples were collected from 

the Frio River by following procedures outlined in the USGS 

“National Field Manual for the Collection of Water-Quality 

Data” (U.S. Geological Survey, variously dated). Prior to 

sampling, field parameters including water temperature, 

dissolved oxygen concentration, pH, turbidity, and SC were 

monitored until they had stabilized and recorded just prior 

to sample collection (Wilde, variously dated). For routine 

samples at the Frio River, stream depths were shallow enough 

to be waded, and samples were collected by using a US DH-81 

with a 1-liter Teflon bottle sampler attached to a wading rod 

as described by Davis (2005). Storm-runoff samples from 

the Frio River and the two other surface-water sites (Culebra 

Creek and Helotes Creek) were collected by using ISCO 3700 

autosamplers (Teledyne Isco, 2018) because these sites are 

not wadable during flash floods. For samples collected by 

using the autosamplers (Frio River and Helotes Creek only), 

flow-weighted stormwater-runoff samples were composited by 

using a point-integrated sampling method. A point-integrated 

sample is a sample accumulated continuously in a sampler that 

is held at a relatively fixed point and that admits a water and 

sediment mixture at a velocity about equal to the instantaneous 

stream velocity at that point (Burkham, 1985). Each sample 

was drawn through a fixed intake mounted at the midpoint in 

the stream channel by using a suction-lift automatic sampler. 

The automatic sampler was programmed to begin sampling 

at a predetermined stage that was selected to ensure that 

streamflow was occurring. The autosampler was programmed 

to collect 500-milliliter samples every 15 minutes (Helotes 

Creek) or 60 minutes (Frio River) and sequentially fill 

each of four 8-liter bottles during the runoff event. The 

difference in sampling intervals between the two sites was 

needed because stormwater-runoff events at Helotes Creek 

are typically of much shorter duration than those at the Frio 

River. The samples were retrieved at the end of each runoff 

event or as soon as practical after all bottles were filled. The 

samples were then placed on ice and transported to the USGS 

South Texas Program Office Laboratory in San Antonio for 

processing. The four 8-liter samples collected by the automatic 

sampler were flow- or stage-weighted and composited into 

a single water-quality sample to represent the event mean 

concentration. Samples collected by wading or autosamplers 

were composited into a Teflon churn. Subsamples for specific 

analysis were withdrawn from the churn and preserved as 

required by the analyzing laboratory. Not all stormwaterrunoff samples, however, could be collected as flow- or stageweighted composites because of variations in rainfall causing 

flows of extended duration, occurrence of multiple discharge 

peaks, mechanical problems with autosamplers, an insufficient 

number of autosamplers for deployment at all of the sites 

prior to a runoff event, or lack of discharge data. In some 

instances, multiple grab samples were collected from the edge 

of the stream or bridge and composited into a single waterquality sample. Compound concentrations from multiple 

grab samples represent instantaneous event concentrations. 

For the purposes of this report, all storm-runoff samples are 

considered to represent instantaneous event concentrations 

and are not used to compute longer term loadings of 

individual compounds.

Routine groundwater samples were collected as 

described in the USGS “National Field Manual for the 

Collection of Water-Quality Data” (U.S. Geological Survey, 

variously dated) at Seco, Parkwood, Shavano, Zarzamora, 

Turtle, and Mission wells, as well as Comal 1 spring, without 

regard to hydrologic conditions (fig. 1; table 1). In addition 

to routine sampling, some samples at unconfined wells 

were collected in response to recharge-generating storms 

and corresponding rises in WLAs. At Seco, Parkwood, 

and Zarzamora wells, a pump was lowered into the well 

for sampling. At Shavano well, a pump was permanently 

installed in situ. Prior to sample collection at Seco, Parkwood, 

Shavano, and Zarzamora wells, three casing volumes were 

purged to ensure that water from the formation was being 

collected. During purging, field parameters including water 

temperature, dissolved oxygen concentration, pH, turbidity, 

and SC were monitored until they had stabilized and were 

recorded just prior to sample collection (Wilde, variously 

dated). At Turtle and Mission wells, samples were collected 

directly at the well head and did not undergo any treatment 

prior to collection. Samples from Comal 1 spring were 

collected by immersing bottles below the water surface 

into or near the spring orifice, avoiding contact with the 

atmosphere and standing surface water. Samples were either 

processed immediately in the field or placed on ice and 

transported to the USGS South Texas Program Office in San 

Antonio for processing. 

Analytical Methods

Most water-quality samples were analyzed for alkalinity, 

major ions, nutrients, pesticides and pesticide degradates, 

and selected environmental isotopes including stable isotopes 

of water (delta deuterium [δD] and δ18O), stable isotopes of 

NO3

 (δ15N-NO3 and δ18O-NO3), and strontium (Sr) isotopes 

(strontium-87 [87Sr]/strontium-86 [86Sr]). Not all samples 

were analyzed for all constituents; specifically, most samples 

from Shavano, Turtle, and Mission wells (the ETN sites) 

were not analyzed for pesticides and pesticide degradates. 

Where analyzed for pesticides, samples from the ETN sites 

were analyzed with a different analytical method. As a result, 

the ETN sites are not considered in the pesticide discussion in 

this report. All laboratory methods used for analysis of waterquality samples are listed in table 2.
12 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Table 2. Measured constituents, laboratory reporting levels, units, method references, and analyzing laboratories for water-quality 

samples collected from sites in the San Antonio segment of the Edwards aquifer, south-central Texas, 2010–16.

[N, nitrogen; mg/L, milligram per liter; NWQL, National Water Quality Laboratory of the U.S. Geological Survey; TKN, total Kjeldahl nitrogen; μg/L, 

microgram per liter; --, not applicable; LRL, laboratory reporting level; P, phosphorus; μS/cm, microsiemens per centimeter at 25 degrees Celsius; MMIL, Metal 

and Metalloid Isotope Laboratory of the U.S. Geological Survey; δ18O, delta oxygen-18; RSIL, Reston Stable Isotope Laboratory of the U.S. Geological Survey; 

δD, deuterium; δ15N-NO3, delta nitrogen-15 in nitrate; δ18O-NO3, delta oxygen-18 in nitrate]

Constituent Laboratory 

reporting level Unit Method reference

Analyzing 

laboratory

Ammonia as N, filtered 0.01 mg/L Fishman, 1993 NWQL

Ammonia plus organic nitrogen as N, filtered 0.07 mg/L Patton and Truitt, 2000 NWQL

Ammonia plus organic nitrogen as N, 

unfiltered (TKN)

0.07 mg/L Patton and Truitt, 2000 NWQL

Bicarbonate, filtered 1 mg/L Rounds, 2006 Analyzed 

in the 

field

Boron, filtered 2.0 μg/L Struzeski and others, 1996 NWQL

Bromide, filtered 0.01 mg/L Fishman and Friedman, 1989 NWQL

Calcium, filtered 0.022 mg/L Fishman, 1993 NWQL

Chloride, filtered 0.02 mg/L Fishman and Friedman, 1989 NWQL

Fluoride, filtered 0.01 mg/L Fishman and Friedman, 1989 NWQL

Magnesium, filtered 0.011 mg/L Fishman, 1993 NWQL

Nitrate plus nitrite as N, filtered 0.01 mg/L Patton and Kryskalla, 2011 NWQL

Nitrite as N, filtered 0.001 mg/L Fishman, 1993 NWQL

pH, unfiltered, lab 0.1 pH unit Fishman, 1993 NWQL

Orthophosphate as P, filtered 0.004 mg/L Fishman, 1993 NWQL

Pesticide and pesticide degradates, filtered Individual LRLs 

provided in Opsahl 

and others (2018)

μg/L Zaugg and others, 1995 NWQL

Phosphorus as P, filtered 0.01 mg/L Patton and Kryskalla, 2003 NWQL

Phosphorus as P, unfiltered 0.01 mg/L Patton and Kryskalla, 2003 NWQL

Potassium, filtered 0.06 mg/L American Water Works Association, 1998 NWQL

Silica, filtered 0.018 mg/L Fishman, 1993 NWQL

Sodium, filtered 0.1 mg/L Fishman, 1993 NWQL

Specific conductance, unfiltered, lab 5 μS/cm Fishman and Friedman, 1989 NWQL

Strontium, filtered 0.2 μg/L Fishman, 1993 NWQL

Strontium-87/strontium-86 ratio, filtered -- ratio Bullen and others, 1996 MMIL

Sulfate, filtered 0.02 mg/L Fishman and Friedman, 1989 NWQL

δ18O in water, unfiltered -- per mil Révész and Coplen, 2008b RSIL

δD in water, unfiltered -- per mil Révész and Coplen, 2008a RSIL

δ15N-NO3

, filtered -- per mil Coplen and others, 2012 RSIL

δ18O-NO3

, filtered -- per mil Coplen and others, 2012 RSIL
Methods 13

Alkalinity was titrated either in the field or at the USGS 

South Texas Program Office by the sampling team. Major 

ions, trace elements, nutrients, pesticides, and pesticide 

degradates were analyzed at the USGS National Water 

Quality Laboratory (NWQL) in Denver, Colorado. Measured 

nutrient species included NO3 plus NO2, NO2, NH4 plus 

dissolved organic N (DON), orthophosphate (ortho-P), 

and phosphorus (P). Calculated nutrient species included 

NO3 (measured NO3 plus NO2 concentration minus the 

measured NO2 concentration), DON (measured NH4 plus 

DON concentration minus the measured NH4 concentration), 

and total dissolved N (sum of the measured concentrations 

of NO3, NO2, NH4, and DON). Dissolved solids were 

calculated as the sum of the measured major ions (calcium, 

magnesium, sodium, potassium, silica, chloride, sulfate, and 

alkalinity [carbonate/bicarbonate expressed as carbonate 

equivalent]), plus the sum of all other measured dissolved 

constituents. Major ions typically make up the bulk of 

the dissolved constituents in a sample (Hem, 1985). All N 

species are reported in units of milligrams per liter as N and 

discussed as N. Ortho-P is reported in units of milligrams 

per liter as P and discussed as P. Results for some pesticides 

and pesticide degradates were qualified as estimated (“E”) 

by the NWQL under laboratory reporting conventions used 

during the study period. A remark code of “E” was used when 

compounds were identified, but concentrations could not 

be accurately quantified for a variety of reasons (Sandstrom 

and others, 2001) or when results are greater than the longterm method detection level and less than the laboratory 

reporting level (LRL) (Childress and others, 1999). The 

87Sr/86Sr samples were analyzed by the USGS Metal and 

Metalloid Isotope Laboratory in Menlo Park, California. The 

δ18O, δD, δ15N-NO3, and δ18O-NO3 samples were analyzed 

by the USGS Reston Stable Isotope Laboratory (RSIL) in 

Reston, Virginia.

Quality Assurance for Water Samples

Quality-assurance (QA) samples including equipment 

and field blanks, split replicate environmental samples, and 

environmental pesticide matrix-spike samples were collected 

and analyzed to evaluate bias and variability of environmental 

samples. Bias is the systematic error inherent in a method 

or caused by some artifact of the measurement system. 

Variability is the extent to which data points in a statistical 

distribution or dataset diverge from the average (or mean) 

value, as well as the extent to which these data points differ 

from each other (Mueller and others, 2015). A total of 75 QA 

samples were collected from rainfall, surface-water, spring, 

and groundwater sites and analyzed for major ions, nutrients, 

trace elements, and pesticide compounds as part of the study 

during 2013–16. All QA sample results collected in the field 

during the study are shown in Opsahl and others (2018). 

QA sample results for water-quality data collected prior to the 

current study are discussed in Opsahl (2012). In addition to 

the QA samples submitted from the field, the USGS NWQL 

analyzed laboratory blanks, replicates, and spikes. 

Blank samples test for bias from introduction of 

contaminants to environmental samples. Two types of blank 

samples were analyzed in this study: equipment and field. 

Equipment blanks are used to demonstrate that samplecollection and sample-processing equipment and equipmentcleaning procedures are not sources of contamination. In 

comparison, field blanks are used to assess contamination 

from field conditions during sampling. All of the blank 

samples were collected, processed, and analyzed in the same 

manner as were the environmental samples. Blanks were 

collected by using high-purity inorganic-free or organicfree water tested by the USGS NWQL and certified to be 

less than the LRL for constituent concentrations. A total of 

25 blanks (combination of inorganic, organic, equipment, 

and field) were collected during this study. Eleven equipment 

blanks (inorganic and [or] organic) were collected from bulk 

rainfall collectors (3 samples), surface-water autosamplers 

(5 samples), groundwater pump equipment (2 samples), and 

pump hoist truck equipment (1 sample) during 2013–16. 

Fourteen field blanks (inorganic or organic) were collected 

from 1 surface-water site (2 samples) and 1 groundwater site 

(12 samples) during 2013–16. 

Major ions, including calcium (Ca), potassium (K), 

chloride (Cl), and fluoride (F), were detected in both the 

equipment blanks and the field blanks (table 3). Ca was a 

commonly detected constituent, occurring at least once in all 

equipment blank type and field blank type samples with the 

exception of the groundwater pump equipment blank sample 

(table 3). The Ca concentrations detected were relatively 

consistent, ranging from 0.024 to 0.077 mg/L for all blank 

types (table 3). Although Ca was detected in most equipment 

and field blanks, the concentrations were several orders of 

magnitude lower than any surface-water or groundwater 

environmental sample concentrations and were assumed to 

not substantially bias the sample results for surface-water 

and groundwater samples. K was only detected in one bulk 

rainfall equipment blank sample. The concentration of K in 

the rainfall equipment blank sample collected from the Urban 

rainfall site 2 on May 6, 2015, was 0.08 mg/L, lower than 

the environmental K concentration of 0.27 mg/L collected on 

May 6, 2015 (Opsahl and others, 2018). The environmental 

K concentrations in samples from this site likely are 

representative of environmental concentrations although about 

a third of the value may be subject to positive bias. Ca was 

detected in all rainfall equipment blanks at concentrations 

similar to those measured in rainfall samples; therefore, Ca 

concentrations in rainfall samples should be considered as 

potentially contaminated and results interpreted with caution.
14 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Table 3. Summary of quality-assurance data for selected compounds detected in equipment blank and field blank water-quality 

samples collected from sites in the San Antonio segment of the Edwards aquifer, south-central Texas, 2013–16.

[LRL, laboratory reporting level; mg/L, milligram per liter; N, nitrogen; results summarized from Opsahl and others (2018)]

Constituent Number of 

samples

Number of 

samples greater 

than LRL

LRL

Minimum sample 

concentration 

greater than LRL

Maximum sample 

concentration 

greater than LRL

Rainfall equipment blank water-quality samples

Calcium, filtered (mg/L) 3 1 0.022 0.024 0.035

Potassium, filtered (mg/L) 3 1 0.03 0.08 0.08

Chloride, filtered (mg/L) 3 3 0.02 0.02 0.06

Fluoride, filtered (mg/L) 3 2 0.01 0.12 0.34

Surface-water autosampler equipment blank water-quality samples

Calcium, filtered (mg/L) 3 2 0.022 0.027 0.077

Chloride, filtered (mg/L) 3 1 0.06 0.06 0.06

Groundwater pump equipment blank water-quality samples

Chloride, filtered (mg/L) 2 1 0.06 0.07 0.07

Pump hoist truck equipment blank water-quality sample

Calcium, filtered (mg/L) 1 1 0.022 0.055 0.055

Ammonia, filtered (mg/L as N) 1 1 0.01 0.01 0.01

Surface-water field blank water-quality sample

Chloride, filtered (mg/L) 1 1 0.02 0.05 0.05

Groundwater field blank water-quality samples

Calcium, filtered (mg/L) 7 1 0.022 0.027 0.027

Chloride, filtered (mg/L) 7 1 0.02 0.03 0.03

Fluoride, filtered (mg/L) 7 1 0.01 0.02 0.02

Ammonia, filtered (mg/L as N) 9 1 0.01 0.02 0.02

Cl was the most commonly detected constituent, 

occurring in the rainfall equipment, surface-water autosampler, 

and groundwater pump equipment blank samples, as well as 

the surface-water field and groundwater field blank samples 

(table 3). The Cl concentrations detected in the blank samples 

ranged from 0.02 to 0.07 mg/L, which was much less than 

the environmental Cl concentrations for all site types (table 3; 

Opsahl and others, 2018). Similar to the results for Ca and 

K, the Cl results from the blank samples are low and are not 

assumed to bias the environmental results. 

F was detected in two rainfall equipment blank samples 

and one groundwater field blank sample (table 3). Both 

rainfall equipment blank samples were collected May 6, 

2015, from different rainfall collectors. The concentrations 

of F detected in the rainfall equipment blank samples were 

0.12 and 0.34 mg/L, which were higher than the associated 

environmental sample concentrations. Concentrations of 

less than (<) 0.01, E0.03, and 0.2 mg/L were measured in 

samples collected from Urban rainfall 2 on May 6, 2015, and 

from the rainfall collector at the Frio River on May 12, 2015, 

and May 13, 2015 (Opsahl and others, 2018). The rainfall 

equipment blank sample F concentrations most likely are 

an artifact of the sample processing and not environmental 

F concentrations. Given the frequency of F detection in 

the rainfall equipment blank samples (two of three) and 

concentrations well above the LRL, all F concentrations 

for rainfall samples were censored from the dataset and 

not used in any data analysis. The concentration of F in the 

groundwater field blank sample collected from the Mission 

well on November 14, 2013, was 0.02 mg/L, whereas the 

environmental F concentration was 0.29 mg/L collected 

from this well on the same date (Opsahl and others, 2018). 

The environmental sample F results for the Mission well 

are representative of environmental concentrations. Given 

that there was only one detection of F in seven groundwater 

field blank samples and that sample had a concentration 

(0.02 mg/L) very close to the LRL (0.01 mg/L), no remaining 

F concentration values were censored from the dataset.
Methods 15

NH4 was detected in the pump hoist truck equipment 

blank sample collected on December 22, 2014, and the 

groundwater field blank sample collected on November 12, 

2013 (table 3 in Opsahl and others, 2018). The NH4

concentration in the pump hoist truck equipment blank sample 

was 0.01 mg/L, and the concentration in the groundwater field 

blank sample was 0.02 mg/L (table 3). These concentrations 

were equal to or higher than the concentrations in the 

associated environmental samples collected from Seco well 

(0.01 mg/L collected on December 23, 2014) and Turtle well 

(0.01 mg/L collected on November 12, 2013) (Opsahl and 

others, 2018), although all concentrations are very low. Given 

that only 2 among all 20 NH4 blank samples in this study had 

a detection of NH4, no data were censored from the study, 

although the possibility that low concentrations of NH4 can 

result from sample processing is acknowledged.

There were no detections of pesticides in any of the 

equipment and field blank samples (Opsahl and others, 

2018). Pesticide concentrations reported from environmental 

samples are assumed to be unaffected by the introduction of 

contaminants originating from sample-collection or sampleprocessing methods.

Replicate samples are used to evaluate variability in 

measurements associated with analytical processes, sampleprocessing protocols, or natural variation and to provide 

information on the reproducibility and precision of sample 

processing and analysis. In this study, split replicate samples 

were used to evaluate sample-processing variability. Precision 

can be evaluated by comparing concentrations between 

environmental and replicate sample pairs and expressed 

as relative percent differences (RPDs). A combination of 

42 split replicate samples (21 replicate pairs) were collected 

and processed from surface-water, spring, and groundwater 

sites during 2013–16 by following the guidelines listed in the 

USGS “National Field Manual for the Collection of WaterQuality Data” (U.S. Geological Survey, variously dated) and 

analyzed for the same constituents used for environmental 

samples including major ions, nutrients, trace elements, 

pesticides, stable isotopes of water and NO3, and strontium 

isotopes (Opsahl and others, 2018). 

RPDs were calculated for each replicate pair of 

constituents having detectable concentrations (summarized 

in table 4). If one or both analyses in a constituent replicate 

pair were reported as less than the applicable LRL, then that 

constituent pair was excluded from the analysis. RPDs were 

calculated by using the following equation:

RPD = +  ( ) ( ) ×  

 C C 1 2 C C 1 2 − / / 2 100 (1)

where

C1 is the constituent concentration, in milligrams 

per liter, from the environmental sample; 

and

C2 is the constituent concentration, in milligrams 

per liter, from the replicate sample.

RPDs were calculated, and results were separated into 

surface-water results or groundwater results (table 4). Twentyfour of thirty-one median RPDs calculated for individual 

constituents from the surface-water replicate samples were less 

than 10 percent (table 4), indicating good reproducibility for 

individual constituents between the environmental and 

replicate surface-water samples. The minimum and maximum 

RPD results calculated for all constituents from surfacewater replicate samples ranged from 0 to 67 percent, with 

a median RPD of 2 percent (table 4). In comparison, 24 of 

25 median RPDs for individual constituents from the 

spring and groundwater replicate samples were less than 

10 percent (table 4), also indicating good reproducibility. 

The RPDs calculated for all constituents from groundwater 

replicate samples ranged from 0 to about 67 percent 

(table 4), with a median RPD of 1 percent. Overall, results 

indicate that there was good to excellent reproducibility 

for most of the constituents. The higher RPDs, generally 

greater than 20 percent, were associated with replicate 

pairs that had lower concentrations that were close in value 

to the LRL (Opsahl and others, 2018). When evaluating 

RPDs, the magnitude of the concentrations being evaluated 

needs to be considered. Although the difference between 

environmental and replicate concentrations may be the same 

for a given constituent, the difference evaluated relative to 

the mean concentration of the replicate pair increases as the 

concentrations being evaluated decrease. Larger RPDs were 

associated with surface-water replicate pairs that had low 

concentrations of bromide, N nutrients (NH4, NO2, organic N, 

and total N), and prometon, as well as for the δ18O of water 

(Opsahl and others, 2018). While the RPDs for replicates 

with lower concentrations often are large, these values do not 

necessarily indicate a lack of laboratory precision. The RPDs 

associated with this study indicated that the sampling and 

analytical procedures for most constituents were consistent 

and reproducible.

Field pesticide matrix-spike samples (hereinafter 

referred to as “pesticide spikes”) are used to assess bias 

and variability from matrix interference or degradation of 

pesticide constituent concentrations that might occur during 

sample processing, storage, and analysis. Pesticide spikes 

are environmental samples to which a known volume and 

concentration of constituents are added; analytical recoveries 

of the spiked constituents are expressed as percentages 

of expected (theoretical) concentrations. Computed 

field-spike recoveries are compared to theoretical and 

laboratory recoveries to evaluate matrix interferences or 

degradation of pesticides.

During 2013–16, pesticide spikes were added to 

8 environmental samples collected from rainfall, surfacewater, spring, and groundwater sites by following guidelines 

listed in the USGS “National Field Manual for the Collection 

of Water-Quality Data” (U.S. Geological Survey, variously 

dated) and analyzed for 83 pesticides and pesticide degradates 

(Opsahl and others, 2018). A spike recovery was calculated 
Table 4. Summary of quality-assurance data of paired environmental and split replicate water-quality samples and associated relative percent diffewater-quality sample pairs collected from surface-water, spring, and groundwater sites in the San Antonio segment of the Edwards aquifer, south-ce[LRL, laboratory reporting level; RPD, relative percent difference in percent (calculated for replicate pairs with concentrations greater than the minimum detectable concentratio*, calculated constituent; --, no data, value not calculated, or concentrations were less than or equal to the LRL; CaCO3, calcium carbonate; SiO2; silicon dioxide; N, nitrogen; P, 

nitrite; μg/L, microgram per liter; E, estimated; DEA, deethylatrazine; CIAT, 2-Chloro-4-isopropylamino-6-amino-s-triazine; δD, delta deuterium in water; δ18O, delta oxygen-1

nitrogen-15 in nitrate; δ18O-NO3, delta oxygen-18 in nitrate; 87Sr, strontium-87; 86Sr, strontium-86; results summarized from Opsahl and others (2018)]

Constituent LRL

Number of replicate sets available 

to calculate for 

surface-water 

samples

Number of 

replicate sets 

greater than LRL 

for surface-water 

samples

Minimum

surface-water RPD 

(percent)

Maximu

surface-wat

(percen

Dissolved solids, filtered, sum of constituents (mg/L)* -- 4 4 1 5 

Hardness, filtered (mg/L as CaCO3)* -- 5 5 1 4 

Calcium, filtered (mg/L) 0.022 5 5 1 4 

Magnesium, filtered (mg/L) 0.011 5 5 1 3 

Potassium, filtered (mg/L) 0.1 5 5 0 5 

Sodium, filtered (mg/L) 0.1 5 5 0 4 

Bromide, filtered (mg/L) 0.01 2 2 2 35 

Chloride, filtered (mg/L) 0.02 5 5 0 1 

Fluoride, filtered (mg/L) 0.01 5 5 0 9 

Silica, filtered (mg/L as SiO2) 0.018 5 5 1 5 

Sulfate, filtered (mg/L) 0.02 5 5 0 0 

Ammonia plus organic nitrogen, filtered (mg/L as N) 0.07 9 8 0 46 

Ammonia plus organic nitrogen, unfiltered (mg/L as N) 0.07 10 10 0 30 

Ammonia, filtered (mg/L as N) 0.01 11 2 0 67 

Nitrate plus nitrite, filtered (mg/L as N) 0.01 11 11 0 6 

Nitrate, filtered (mg/L as N)* -- 11 11 0 4 

Nitrite, filtered (mg/L as N) 0.001 11 8 0 40 

Organic nitrogen, filtered (mg/L as N)* 0.02–0.18 10 1 0 0 

Organic nitrogen, unfiltered (mg/L as N)* 0.07 10 2 9 31 

Orthophosphate, filtered (mg/L as P) 0.004 5 2 0 0 

Phosphorus, filtered (mg/L as P) 0.01 4 0 -- -- 

Phosphorus, unfiltered (mg/L as P) 0.01 4 0 -- -- 

Total nitrogen [NO3 + NO2 + NH3 + organic-N],

filtered (mg/L)

0.14–1.9 9 8 0 13 

Total nitrogen [NO3 + NO2 + NH3 + organic-N],

unfiltered (mg/L)

1.2–2.6 10 10 0 16 
Constituent LRL

Number of replicate sets available 

to calculate for 

surface-water 

samples

Number of 

replicate sets 

greater than LRL 

for surface-water 

samples

Minimum

surface-water RPD 

(percent)

Maximu

surface-wat

(percen

Strontium, filtered (μg/L) 0.2 9 9 0 2 

Boron, filtered (μg/L) 2.0 4 4 0 5 

DEA (CIAT), filtered, recoverable (μg/L) 0.0 2 0 -- -- 

Atrazine, filtered, recoverable (μg/L) 0.008 2 0 -- -- 

Desulfinyl-fipronil, filtered, recoverable (μg/L) 0.012 2 1 0 0 

Prometon, filtered, recoverable (μg/L) 0.012 2 1 15 15 

Simazine, filtered, recoverable (μg/L) 0.006 2 0 -- -- 

δD in water, unfiltered, per mil -- 2 2 1 8 

δ18O in water, unfiltered, per mil -- 2 2 4 22 

δ15N-NO3, filtered, per mil -- 2 1 2 2 

δ18O-NO3, filtered, per mil -- 2 1 6 6 

87Sr/86Sr, filtered, ratio -- 2 2 0 0 
Table 4. Summary of quality-assurance data of paired environmental and split replicate water-quality samples and associated relative percent diffewater-quality sample pairs collected from surface-water, spring, and groundwater sites in the San Antonio segment of the Edwards aquifer, south-ceContinued [LRL, laboratory reporting level; RPD, relative percent difference in percent (calculated for replicate pairs with concentrations greater than the minimum detectable concentratio*, calculated constituent; --, no data, value not calculated, or concentrations were less than or equal to the LRL; CaCO3, calcium carbonate; SiO2; silicon dioxide; N, nitrogen; P, nitrite; μg/L, microgram per liter; E, estimated; DEA, deethylatrazine; CIAT, 2-Chloro-4-isopropylamino-6-amino-s-triazine; δD, delta deuterium in water; δ18O, delta oxygen-1nitrogen-15 in nitrate; δ18O-NO3, delta oxygen-18 in nitrate; 87Sr, strontium-87; 86Sr, strontium-86; results summarized from Opsahl and others (2018)] Constituent Median surface-water concentration Number of repli- cate sets avail- able to calculate for groundwater samples2 Number of replicate sets greater than LRL for groundwater samples Minimum groundwater RPD (percent) Maximum groundwater RPD (percent) M

grou

(pe

Dissolved solids, filtered, sum of constituents (mg/L)* 256 8 8 0 2 Hardness, filtered (mg/L as CaCO3)* 235 8 8 0 6 Calcium, filtered (mg/L) 63.55 8 8 0 8 Magnesium, filtered (mg/L) 17.55 8 8 0 2 Potassium, filtered (mg/L) 1 8 8 1 12 Sodium, filtered (mg/L) 8.76 8 8 0 5 Bromide, filtered (mg/L) 0.1 8 8 0 5 Chloride, filtered (mg/L) 13.2 8 8 0 1 Fluoride, filtered (mg/L) 0.13 8 8 0 11 Silica, filtered (mg/L as SiO2) 12.3 8 8 0 7 Sulfate, filtered (mg/L) 35.8 8 8 0 1 Ammonia plus organic nitrogen, filtered (mg/L as N) 0.14 2 0 -- -- Ammonia plus organic nitrogen, unfiltered (mg/L as N) 0.17 4 0 -- -- Ammonia, filtered (mg/L as N) 0.01 8 0 -- -- Nitrate plus nitrite, filtered (mg/L as N) 1.05 9 9 0 3 Nitrate, filtered (mg/L as N)* 1.05 8 8 0 0 Nitrite, filtered (mg/L as N) 0.004 8 0 -- -- Organic nitrogen, filtered (mg/L as N)* 0.13 5 0 -- -- Organic nitrogen, unfiltered (mg/L as N)* 0.12 4 0 -- -- Orthophosphate, filtered (mg/L as P) 0.005 8 8 0 13 Phosphorus, filtered (mg/L as P) -- 4 0 -- -- Phosphorus, unfiltered (mg/L as P) -- 4 2 0 67 Total nitrogen [NO3 + NO2 + NH3 + organic-N],

filtered (mg/L)

1.65 2 0 -- -- 

Total nitrogen [NO3 + NO2 + NH3 + organic-N], 

unfiltered (mg/L)

1.35 4 0 -- -- 
Constituent

Median 

surface-water 

concentration

Number of replicate sets available to calculate 

for groundwater 

samples2

Number of 

replicate sets 

greater than LRL 

for groundwater 

samples

Minimum 

groundwater 

RPD 

(percent)

Maximum 

groundwater 

RPD 

(percent)

M

grou

(pe

Strontium, filtered (μg/L) 395 4 4 1 3 Boron, filtered (μg/L) 62.0 8 8 0 3 DEA (CIAT), filtered, recoverable (μg/L) E0.009 4 2 0 5 Atrazine, filtered, recoverable (μg/L) -- 4 2 0 0 Desulfinyl-fipronil, filtered, recoverable (μg/L) 0.002 4 0 -- -- Prometon, filtered, recoverable (μg/L) 0.021 4 0 -- -- Simazine, filtered, recoverable (μg/L) -- 4 3 0 0 δD in water, unfiltered, per mil −14.85 4 4 0 2 δ18O in water, unfiltered, per mil −1.93 4 4 0 2 δ15N-NO3, filtered, per mil 6.28 4 4 0 3 δ18O-NO3, filtered, per mil 2.15 4 4 2 7 87Sr/86Sr, filtered, ratio 0.70773 6 6 0 0 

1Relative percent differences were not calculated when estimated concentrations (E) were reported for one or both replicate samples.2Samples from the Comal 1 spring site at Comal Springs (fig. 1) were for the purpose of this report considered groundwater samples representative of deeper, confined ground
20 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

for constituents with environmental concentrations greater than 

the LRL. For each constituent with a censored value (value 

less than the LRL) in the environmental sample, a range of 

spike recovery was calculated by using zero (0) for the lowest 

possible concentration and the LRL for the specified constituent 

for the highest possible concentration. An average recovery for 

censored values less than the LRL then was computed by using 

the low recovery and the high recovery values. Negative values 

for minimum percent recovery are possible when the measured 

result of the spike sample is less than the LRL. The percent 

recovery of the spiked pesticide concentrations was calculated 

by using the following equations:

Percent recovery = × s u expd ([C C− ] ) 100 / ( ) C (2)

where

Cs is the measured result of the spiked sample, 

in micrograms per liter, from the 

environmental sample;

Cu is the measured result of the unspiked 

sample, in micrograms per liter, from the 

environmental sample; and

Cexpd is the expected or theoretical concentration 

added to the spiked sample, in micrograms 

per liter, which is calculated by using the 

following equation:

C C expd = × sol s V V pike sampl ( ) / ( ) (3)

where

Csol is the concentration of spike solution, in 

micrograms per liter;

Vspike is the amount of spike added, in milliliters; and

Vsampl is the spiked sample volume, in liters. 

Minimum, maximum, and median percent recoveries 

of each compound were computed for surface-water, spring, 

and groundwater sites and also individually for the single 

rainfall sample (Opsahl and others, 2018). Spike recoveries 

higher than 100 percent and outside the control limits of the 

laboratory indicate a positive bias for a particular constituent; 

the actual concentrations probably are lower than the 

reported concentrations. Pesticide spike recoveries lower than 

100 percent indicate a negative bias for a particular constituent; 

the actual concentrations probably are higher than the reported 

concentrations for the specified constituent. Pesticide spike 

recoveries in the range of 85–115 percent generally indicate that 

the reported environmental pesticide concentration is a close 

approximation of the actual pesticide concentration and not 

greatly affected by matrix influence or degradation.

Overall, the single rainfall pesticide spike sample had the 

best recovery for all constituents (median = 87 percent; n = 1) as 

compared to surface-water spike samples (median = 65 percent; 

n = 3) and spring and groundwater samples (median = 68 

percent; n = 4) (Opsahl and others, 2018). The most commonly 

detected pesticides in environmental samples collected during 

2013–16 were atrazine, DEA, prometon, and simazine. Pesticide 

spike recoveries for atrazine were highest in the rainfall spike 

sample (95 percent) and the surface-water spike samples 

(minimum RPD = 70 percent; maximum RPD = 102 percent; 

median RPD = 90 percent; n = 3) and lowest in the spring 

and groundwater spike samples (minimum RPD = 76 percent; 

maximum RPD = 96 percent; median RPD = 83 percent). These 

spike recoveries indicate that the reported concentrations of 

atrazine in the rainfall samples are close in value to the actual 

concentrations of atrazine, whereas the reported concentrations 

of atrazine in surface-water, spring, and groundwater samples 

possibly are less than the actual concentrations. Percent recovery 

of DEA in rainfall samples was 71 percent (Opsahl and others, 

2018). Lower DEA recoveries were observed in the surfacewater samples (minimum RPD = 49 percent; maximum RPD 

= 78 percent; median RPD = 59 percent) and the spring and 

groundwater samples (minimum RPD = 49 percent; maximum 

RPD = 63 percent; median RPD = 61 percent). Consistent with 

results from Martin and others (2009), reported concentrations of 

DEA are possibly lower than the actual concentrations. Percent 

recovery of prometon in rainfall samples was about 60 percent. 

Similar prometon recoveries were observed in the surfacewater samples (minimum RPD = 62 percent; maximum RPD 

= 74 percent; median RPD = 66 percent) and the spring and 

groundwater samples (minimum RPD = 42 percent; maximum 

RPD = 77 percent; median RPD = 60 percent). These results 

indicate that reported concentrations of prometon are possibly 

lower than the actual concentrations. Median percent recovery 

of simazine in rainfall samples was about 103 percent. Lower 

simazine recoveries were observed in the surface-water samples 

(minimum RPD = 81 percent; maximum RPD = 110 percent; 

median RPD = 83 percent) and the spring and groundwater 

samples (minimum RPD = 74 percent; maximum RPD = 85 

percent; median RPD = 78 percent). The percent recoveries for 

simazine in surface-water and spring and groundwater samples 

indicate that the reported concentrations are possibly lower than 

the actual concentrations.

Climate Conditions

The study area is generally prone to climatic and 

hydrologic extremes (Griffiths and Strauss, 1985; Jones, 

1991). The climatic and hydrologic conditions during the 

study period varied considerably (fig. 3). The study began 

in July 2013, when the region was undergoing drought 

conditions. From July 2013 through October 2014, rainfall 

in the study area was generally low and generally consistent 

with drought conditions. The Palmer Drought Severity Index 

(PDSI) is an index calculated from climatological data that is 

used to estimate dryness associated with drought conditions; 

the PDSI is particularly useful for estimating cumulative effects 

over monthly time scales (National Oceanic and Atmospheric 

Administration, 2017c). Indicators of climatic and hydrologic 
Methods 21

Jan. 2017

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Palmer Drought Severity Index, 

Edwards Plateau 

Daily rainfall at Rainfall

data 3 site3, in inches

Daily mean spring discharge at Comal Springs3, 

in cubic feet per second 

620

630

640

650

660

670

680

690

700

-5

-4

-3

-2

-1

0

1

2

3

4

5

0

2

4

6

0

50

100

150

200

250

300

350

400

450

500

Daily median water-level altitude at J–17 well3

,

in feet above National Geodetic Vertical

Datum of 1929

A

B

C

Dry period1 Wet period2

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the ”wet period.”

3

Short name (fig. 1; table 1).

Figure 3. Time series (July 2013–December 2016) of A, Palmer Drought Severity Index for the Texas Edwards 

Plateau (National Oceanic and Atmospheric Administration, 2017c); B, spring discharge for Comal Springs

(U.S. Geological Survey, 2017) and water-level altitude for J-17 well (Edwards Aquifer Authority, 2017b); and 

C, daily rainfall recorded at Rainfall data 3 (National Oceanic and Atmospheric Administration, 2017d). 
22 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

conditions including spring discharge at Comal Springs 

and the WLA at J-17 well were consistent with the PDSI, 

indicating that drought conditions and relatively little 

aquifer recharge were occurring during the early part of the 

study. In November 2014 and again in April and May 2015, 

widespread rains fell regionally including in the contributing 

zone and recharge zone of the aquifer. Substantial changes 

followed over several months, including an increased spring 

discharge, a rise in WLA at J-17 well, and a shift in the PDSI 

to normal and then moist conditions. Relatively dry conditions 

in summer and fall 2015 were accompanied by a moderate 

decline in the PDSI, J-17 well WLAs, and Comal Springs 

discharge, but not a return to drought conditions. Winter rains 

in late 2015 and early 2016 initiated another prolonged rise 

in the PDSI, J-17 well WLAs, and Comal Springs discharge. 

In May and June 2016 heavy rains contributed further to this 

prolonged period of active aquifer recharge. At the conclusion 

of the study in December 2016, the WLA at J-17 well had 

risen about 61.6 ft from its lowest point on September 3, 

2014 (WLA = 625.9 ft), to its highest point on December 11, 

2016 (WLA = 687.5 ft). The determination of dry and wet 

periods was based not only on the PDSI but also on hydrologic 

conditions including WLAs at J-17 well and Comal Springs 

discharge. For purposes of this report, the period between 

July 2013 and May 2015 is referred to as the “dry period,” and 

June 2015 through December 2016 is referred to as the “wet 

period” (fig. 3).

Hydrologic Conditions

Surface Water

The Frio River, a stream that flows across the western 

rural portion of the recharge zone, has streamflow that is 

usually sustained by spring discharge, making this site the 

most amenable of the study’s surface-water sites to continuous 

monitoring. Streamflow patterns and water-quality samplecollection dates for the Frio River showed long periods of 

base flow that were interrupted by rapid streamflow increases 

following rainfall and runoff in the contributing zone (fig. 4). 

Daily median streamflow in the Frio River was at base flow 

(less than 40 cubic feet per second [ft3/s]) at the onset of the 

study and remained at base flow until May 2014. The largest 

storm during the study occurred in May 2015, resulting in a 

daily median streamflow reaching greater than 7,100 ft3/s. 

SC values during base flow ranged from about 420 to 

520 microsiemens per centimeter at 25 degrees Celsius. 

During stormwater-runoff events, SC values decreased in 

response to storm pulses that were dominated by runoff from 

recent rainfall. During some stormwater-runoff events (for 

example, July 2014), however, a pulse of relatively high 

SC water occurred as the stormwater-runoff peak receded. 

Identifying the source of this higher SC was not in the 

scope of this study, though it might represent a flush from 

the shallow aquifer of higher conductance water that had 

accumulated between storm events.

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Date

Daily median specific conductance, in microsiemens per centimeter

at 25 degrees Celsius

Daily median discharge, in cubic feet per second

5

50

500

5,000

10,000

280

330

380

430

480

530

580

630

680

Frio River (fig. 1; table 1)

Daily median specific

conductance

Daily median discharge

Water-quality sample

EXPLANATION

Figure 4. Daily median 

discharge, specific conductance, 

and sample-collection dates at 

Frio River (station 0819500), San 

Antonio segment of the Edwards 

aquifer, south-central Texas, 

July 2013–December 2016.
Hydrologic Conditions 23

Groundwater

Continuous WLA and SC data collected at the different 

groundwater sites showed different responses to regional 

climatic and hydrologic conditions that appeared to vary on 

the basis of individual site properties (fig. 5A–G). WLA in 

the wells and discharge at Comal Springs all had notable 

increases following the transition from the dry to wet period. 

Among the three unconfined wells (Seco, Parkwood, and 

Shavano wells), Parkwood and Shavano wells were more 

immediately responsive to recharge events relative to the 

deeper Seco well. Similarly, in comparison to Seco well, SC 

at Parkwood and Shavano wells changed substantially in 

concert with changes in WLA in response to recharge events. 

These results suggest that dilute surface-water recharge more 

directly affects WLA and water chemistry at Parkwood and 

Shavano wells relative to Seco well. SC at Seco well was 

notably lower than at Parkwood or Shavano wells, indicating 

more dilute recharge sources. Over the entire period of 

study, the patterns of change in WLA at the unconfined 

wells were similar, although the net change at Shavano well 

(about 41 ft) was less than at Seco (about 48 ft) and Parkwood 

(about 52 ft) wells.

Turtle and Mission wells are relatively deep confined 

wells located downgradient of the unconfined zone and 

the unconfined wells (fig. 1). WLA was not measured at 

Turtle and Mission wells because Turtle well was in use 

as a production well and the Mission well is artesian. 

Alternatively, WLA at J-17 well (fig. 3), which is used to 

represent regional hydrologic conditions, is hereinafter used 

to provide context for Turtle and Mission wells (fig. 5E, 

F) and to facilitate comparisons with other sites. Among 

the three confined wells (fig. 5D–F), little change in SC 

occurred relative to the unconfined wells (fig. 5A–C), 

regardless of hydrologic conditions. Somewhat larger 

fluctuations of SC occurred at the Mission well, although 

these changes do not generally occur relative to changes in 

hydrologic conditions. The Mission well is located very near 

the freshwater/saline-water interface (fig. 1) in a well field 

of multiple wells; the changes in SC at the Mission well 

are hypothesized to reflect mixing with small and variable 

amounts of downdip saline water in response to well-field 

dynamics or other unknown factors as discussed by Musgrove 

and others (2010). Similar to the deeper, confined wells, 

Comal 1 spring showed relatively small changes in SC over 

the study period (fig. 5G). The general lack of variability in 

SC at the confined groundwater sites indicates that individual 

recharge events do not directly affect water chemistry at these 

sites, even with their large changes in WLA (or discharge 

for Comal 1 spring) in response to changes in hydrologic 

conditions. It is likely that flow paths to these wells are 

sufficiently long for mixing processes to homogenize 

groundwater chemistry. 
24 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Date

Daily median water-level

altitude, in feet

above NAVD 88

700

720

740

760

780

500

560

620

680

740

Daily median specific

conductance, in

μS/cm at 25°C

460

520

580

640

700

Daily median specific

conductance, in

μS/cm at 25°C

Daily median water-level

altitude, in feet

above NAVD 88

630

650

670

690

710

Daily median specific

conductance, in microsiemens

per centimeter at 25 degrees

Celsius (μS/cm at 25°C)

340

400

460

520

580

420

480

540

600

660

Daily median specific

conductance, in

μS/cm at 25°C

760

780

800

820

840

Daily median water-level

altitude, in feet above

North American Vertical

Datum of 1988 (NAVD 88)

670

690

710

730

750

Daily median water-level

altitude, in feet

above NAVD 88

A. Seco well (fig. 1; table 1)

B. Parkwood well (fig. 1; table 1)

C. Shavano well (fig. 1; table 1)

D. Zarzamora well (fig. 1; table 1)

Daily median specific

conductance

Daily median

water-level altitude

Water-quality sample

EXPLANATION

Daily median specific

conductance

Daily median

water-level altitude

Water-quality sample

EXPLANATION

Daily median specific

conductance

Daily median

water-level altitude

Water-quality sample

EXPLANATION

Daily median specific

conductance

Daily median

water-level altitude

Water-quality sample

EXPLANATION

Figure 5. Water-level altitude, specific conductance, and sample-collection times at A, Seco Well, B, Parkwood well, C, Shavano well, 

D, Zarzamora well, E, Turtle well, F, Mission well, and G, Comal 1 spring in the San Antonio segment of the Edwards aquifer, southcentral Texas, July 2013–December 2016. Water-level altitude at J-17 well was used to represent regional hydrologic conditions for 

Turtle and Mission wells (E and F, respectively).
Hydrologic Conditions 25

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Date

400

460

520

580

640

460

520

580

640

700

Daily median specific

conductance, in

microsiemens per

centimeter at 25 degrees

Celsius (μS/cm at 25°C)

Daily median specific

conductance, in

μS/cm at 25°C

Daily median specific

conductance, in

μS/cm at 25°C

Daily median water-level

altitude of J–17 well,

in feet above NGVD 29

Daily median water-level

altitude of J–17 well, in feet

above National Geodetic

Vertical Datum of 1929 (NGVD 29)

620

640

660

680

700

620

640

660

680

700

480

540

600

660

720

50

150

250

350

450

Comal Springs

daily mean discharge,

in cubic feet per second

F. Mission well2 (fig. 1; table 1)

E. Turtle well1

 (fig. 1; table 1)

G. Comal 1 spring3

 (fig. 1; table 1)

1

Daily median specific conductance is for Turtle well. Daily median water-level altitude is for J–17 well (fig. 1; table 1).

2

Daily median specific conductance is for Mission well. Daily median water-level altitude is for J–17 well (fig. 1; table 1).

3

Daily median specific conductance is for Comal 1 spring. Daily mean spring discharge is for Comal Springs (fig. 1; table 1).

Daily median specific

conductance

Daily mean discharge

Water-quality sample

EXPLANATION

Daily median specific

conductance

Daily median

water-level altitude

Water-quality sample

EXPLANATION

Daily median specific

conductance

Daily median

water-level altitude

Water-quality sample

EXPLANATION

Figure 5. Water-level altitude, specific conductance, and sample-collection times at A, Seco Well, B, Parkwood well, C, Shavano well, 

D, Zarzamora well, E, Turtle well, F, Mission well, and G, Comal 1 spring in the San Antonio segment of the Edwards aquifer, southcentral Texas, July 2013–December 2016. Water-level altitude at J-17 well was used to represent regional hydrologic conditions for 

Turtle and Mission wells (E and F, respectively).—Continued
26 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Geochemical Conditions

Meteoric Water Line

Stable isotopes of hydrogen (δD) and oxygen (δ18O) 

are commonly used in studies of hydrologic processes to 

provide insight into numerous factors such as groundwater 

origin and recharge sources, mixing processes, evaporation, 

and response of karst systems to storms (Craig, 1961; Gat, 

1981; Lakey and Krothe, 1996; Jones and others, 2000; 

Desmarais and Rojstaczer, 2002; Maloszewski and others, 

2002). δD and δ18O values for all samples collected during 

this study (rainfall, surface water, and groundwater) have a 

large range, from −67.5 to −1.4 per mil and −9.87 to −1.15 per 

mil, respectively (fig. 6A, B, C). The large range in values 

is primarily accounted for by differences in rainfall samples 

(fig. 6A) and surface-water samples (fig. 6B); groundwater 

samples have a much narrower range of isotopic values 

(fig. 6C, D). A comparison of δD and δ18O data with the global 

meteoric water line (GMWL) (fig. 6D), which represents the 

relation for the isotopic composition of rainfall around the 

globe (Craig, 1961), and a local meteoric water line (LMWL) 

(Pape and others, 2010) indicates that all samples are meteoric 

in origin. The meteoric signature for groundwater samples 

indicates that recharging surface water and groundwater are 

not affected by processes such as evaporation or extensive 

water-rock interaction that would offset samples from the 

GMWL or LMWL. 

Variability of δD and δ18O values among rainfall, surfacewater, and groundwater samples can provide insight into the 

susceptibility of different zones of the aquifer to inputs of 

nutrients and pesticides. A narrower range of δD and δ18O 

values was measured for groundwater (fig. 6C) relative to 

rainfall (fig. 6A) and recharging surface water (fig. 6B), 

consistent with mixing and homogenization of recharge 

within the aquifer. The δD and δ18O values from the deeper, 

confined wells (Zarzamora, Turtle, and Mission wells) cluster 

tightly around median groundwater δD and δ18O values of 

−23.3 and −4.17 per mil, respectively (fig. 6D). δD and δ18O 

values measured in samples from the shallow, unconfined 

wells are more variable than those from the deeper, confined 

wells (fig. 6D). The narrower range of δD and δ18O values 

measured for the confined groundwater samples relative 

to the range of δD and δ18O measured for the unconfined 

groundwater samples is consistent with the characterization of 

confined groundwater as a deeper well-mixed, homogeneous 

endmember in the aquifer (Musgrove and others, 2010). 

Several samples from the unconfined recharge zone wells 

(specifically, those collected from Shavano and Parkwood 

wells) have lower stable-isotope values relative to most of the 

groundwater samples (fig. 6D). The samples with relatively 

low isotope values were usually collected at recharge zone 

wells during wet conditions, often following a recharge 

event. Rainfall, and corresponding recharge, associated with 

tropical cyclonic storms has been shown to have isotopically 

lower stable-isotope values (Lawrence and Gedzelman, 1996; 

Lawrence, 1998). Additionally, rainfall amount has been 

shown to be an important control on stable-isotope values for 

rainfall in central Texas, with lower isotopic values generally 

associated with increasing rain amounts (Pape and others, 

2010). Thus, the isotopic composition of samples with lower 

isotope values from unconfined wells is consistent with 

mixing of ambient groundwater with a component derived 

from a recent, isotopically light recharge source and provides 

evidence that such areas are more susceptible to contamination 

and would be more vulnerable if a source of a compound of 

concern is present.

Water-Rock Interaction

Geochemical evolution processes that occur as a result 

of interaction between water and rock in carbonate aquifers 

also provide insight into groundwater residence time, flow 

paths, diffuse and conduit recharge sources, mineral-solution 

reactions, and mixing (Moore and others, 2009; Musgrove 

and others, 2010). Magnesium (Mg)/Ca and Sr/Ca ratios in 

carbonate groundwater typically increase along flow paths as 

a result of progressive water-rock interaction (Plummer, 1977; 

Trudgill and others, 1980; Lohmann, 1988; Fairchild and 

others, 2000; Musgrove and Banner, 2004). 87Sr/86Sr values 

in the Edwards aquifer generally decrease with increasing 

amounts of water-rock interaction, approaching values similar 

to those of the Cretaceous-age limestone rocks, which have 

87Sr/86Sr values ranging from 0.7074 to 0.7077 (Koepnick and 

others, 1985; Oetting and others, 1996). Longer groundwater 

residence times provide increased opportunity for mineralsolution reactions in Edwards aquifer groundwater that tend 

to result in higher Mg/Ca and Sr/Ca ratios and lower Sr 

isotope compositions. 

The relation between Sr isotopes and Mg/Ca ratios 

(fig. 7A) for this study is consistent with regional patterns 

described by Musgrove and others (2010), where Mg/Ca 

ratios increase as 87Sr/86Sr values decrease from the shallow, 

unconfined groundwater samples to deeper, confined 

groundwater samples. This pattern indicates increasing 

water-rock interaction and mineral-solution reactions with 

the carbonate aquifer rocks controlling fluid compositions as 

groundwater evolves from shallow, unconfined groundwater 

to deeper, confined groundwater. These consistent changes 

result from increased residence time and longer groundwater 

flow paths associated with the deeper, confined groundwater 

sites (Zarzamora well, Turtle well, Mission well, and Comal 

1 spring). In contrast, the higher 87Sr/86Sr values in shallow, 

unconfined groundwater sites (Parkwood, Shavano, and Seco 

wells) indicate shorter residence times and shorter flow paths. 

Additionally, 87Sr/86Sr values are generally more temporally 

variable at the shallow, unconfined wells, likely as a result of 

groundwater mixing with recent, more geochemically variable 

recharge, particularly during the wet period. 
Delta oxygen-18, in per mil

–20 –15 –10 –5 0 5

Delta oxygen-18, in per mil

–20 –15 –10 –5 0 5

Delta oxygen-18, in per mil

–20 –15 –10 –5 0 5

Delta oxygen-18, in per mil

–6.5 –6.0 –5.5 –5.0 –4.5 –4.0 –3.5

Delta deuterium, in per mil

–140–120–100

–80–60–40–20

0

20

Delta deuterium, in per mil

–140–120–100

–80–60–40–20

0

20

Delta deuterium, in per mil

–140–120–100

–80–60–40–20

0

20

Delta deuterium, in per mil

–34–32–30–28–26–24–22–20–18–16

Global Meteoric Water Line Local Meteoric Water Line Local Meteoric Water Line Global Meteoric Water Line

Local Meteoric Water Line Global Meteoric Water Line Local Meteoric Water Line Global Meteoric Water Line

Area enlarged on D

SMALLERSCALE

A B

C D

Figure 6. Relation between deuterium and oxygen isotopes measured in A, rainfall samples, B, surface-water samples, and C and D, groundwater sSan Antonio segment of the Edwards aquifer, south-central Texas, 2013–16. C and D show the same groundwater samples at different scales. Local (global (Craig, 1961) meteoric water lines are shown for comparison.
28 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Date

Strontium-87 to strontium-86 isotopic ratio

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

0.7076

0.7078

0.7080

0.7082

0.7084

0.7086

0.7088

Strontium-87 to strontium-86 isotopic ratio Magnesium to calcium

molarratio

0

0.10

0.20

0.30

0.40

0.50

0.60

0.7076

0.7078

0.7080

0.7082

0.7084

0.7086

0.7088

A

B

Groundwater sample, by

site (fig. 1; table 1)—

Dashed where gap in

data

 Seco well

 Parkwood well

 Shavano well

 Zarzamora well

 Turtle well

 Mission well

 Comal 1 spring

EXPLANATION

Groundwater sample, by

site (fig. 1; table 1)

 Seco well

 Parkwood well

 Shavano well

 Zarzamora well

 Turtle well

 Mission well

 Comal 1 spring

EXPLANATION

Figure 7. A, Relation between strontium-87/strontium-86 isotopic ratio and magnesium to calcium molar ratio and B, time series of 

strontium-87/strontium-86 isotopic ratio measured in groundwater samples collected from the San Antonio segment of the Edwards 

aquifer, south-central Texas, 2013–16.
Geochemical Conditions 29

Major-Ion Distributions

Rainfall

In the Edwards aquifer, major-ion geochemistry has 

been used to characterize spatial and temporal variability 

in water quality (Musgrove and others, 2010), although 

the role of rainfall and corresponding aquifer recharge in 

understanding major-ion dynamics is less well documented. 

Major-ion concentrations in rainfall samples collected from 

both the rural and urban rainfall sites were generally low 

and, in some cases, less than the LRL (table 5). Median 

concentrations were generally similar between rural and 

urban rainfall samples, but there were only four rural rainfall 

samples, which limits comparison for differences between 

rural and urban rainfall. 

Surface Water

Major-ion geochemistry in surface-water samples was 

variable, with considerable differences in ranges and median 

concentrations among sites (table 6). These differences likely 

result from differences in hydrologic conditions at the time of 

sample collection. Samples from the Frio River were primarily 

collected during base-flow conditions, whereas samples 

from the other surface-water sites were collected following 

storms. Stormwater runoff is diluted by rainfall, which has low 

concentrations of major ions (table 5). Samples from the Frio 

River collected during base flow had higher concentrations of 

Ca, Mg, and Sr than did the other surface-water sites (table 6), 

which likely reflect groundwater contributions to base flow 

and lack of dilution with rainfall. Concentrations of other ions, 

such as Cl and sulfate (SO4), at Seco Creek and Culebra Creek 

were notably lower than concentrations at either Helotes Creek 

or the Frio River; however, the small sample size for Seco 

Creek (n = 2) and Culebra Creek (n = 3) limits comparison 

among surface-water sites.

Table 5. Summary of selected major ions measured in rainfall samples collected from Frio River and Urban rainfall 1, 2, and 3, San 

Antonio segment of the Edwards aquifer, south-central Texas, 2012–16.

[LRL, laboratory reporting level; Ca, calcium; mg/L, milligram per liter; Mg, magnesium; Na, sodium; HCO3, bicarbonate; ND, not determined; Cl, chloride; 

SO4, sulfate; Sr, strontium; μg/L, microgram per liter; <, less than]

Map identifier 

(fig. 1; table 1)

Short name 

(fig. 1; table 1) Constituent Number of samples

Number of 

concentrations 

equal to or 

greater than 

the LRL

Minimum 

value

Maximum 

value

Median 

value

7 Frio River1 Ca (mg/L) 4 4 0.07 0.99 0.15

Mg (mg/L) 4 4 0.02 0.05 0.04

Na (mg/L) 4 4 0.14 0.26 0.22

HCO3

 (mg/L) ND ND ND ND ND

Cl (mg/L) 4 4 0.32 0.53 0.39

SO4

 (mg/L) 4 4 0.3 1.61 0.81

Sr (μg/L) 4 3 0.5 2.5 1.0

4, 5, 6 Urban rainfall 

1, 2, 3

Ca (mg/L) 9 9 0.03 1.68 0.47

Mg (mg/L) 9 7 0.02 0.15 0.05

Na (mg/L) 9 8 <0.06 0.83 0.23

HCO3

 (mg/L) 2 1 <1 2 ND

Cl (mg/L) 9 9 0.07 1.84 0.45

SO4

 (mg/L) 9 9 0.27 5.06 1.05

Sr (μg/L) 9 8 0.2 6.2 1.5

1

Rainfall samples collected at U.S. Geological Survey streamflow-gaging station 08195000 Frio River at Concan, Texas (Frio River site), were considered 

representative of rural rainfall. 
30 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Table 6. Summary of selected major ions measured in water samples collected from Frio River, Seco Creek, Culebra Creek, and 

Helotes Creek, San Antonio segment of the Edwards aquifer, south-central Texas, 2010–16.

[LRL, laboratory reporting level; Ca, calcium; mg/L, milligram per liter; Mg, magnesium; Na, sodium; HCO3, bicarbonate; Cl, chloride; SO4, sulfate; Sr, 

strontium; μg/L, microgram per liter; ND, not determined]

Map identifier 

(fig. 1; table 1)

Short name

(fig. 1; table 1) Constituent Number of samples

Number of 

concentrations 

equal to or 

greater than 

the LRL

Minimum 

value

Maximum 

value

Median 

value

7 Frio River Ca (mg/L) 46 46 44.2 87.1 68.1

Mg (mg/L) 46 46 6.24 18.7 16.4

Na (mg/L) 46 46 2.98 9.72 8.20

HCO3 (mg/L) 38 38 156 351 240

Cl (mg/L) 46 46 4.73 15.8 12.9

SO4 (mg/L) 46 46 9.14 60.5 28.8

Sr (μg/L) 30 30 140 461 362

8 Seco Creek Ca (mg/L) 2 2 28.5 32.6 30.1

Mg (mg/L) 2 2 1.88 3.99 3.99

Na (mg/L) 2 2 1.13 1.43 1.28

HCO3 (mg/L) 0 ND ND ND ND

Cl (mg/L) 2 2 1.9 2.24 2.07

SO4 (mg/L) 2 2 5.99 3.66 4.83

Sr (μg/L) 2 2 53.9 61.0 57.5

9 Culebra Creek Ca (mg/L) 3 3 24.4 30.8 30.7

Mg (mg/L) 3 3 1.22 1.87 2.13

Na (mg/L) 3 3 1.05 1.62 1.47

HCO3 (mg/L) 0 ND ND ND ND

Cl (mg/L) 3 3 0.92 3.70 1.35

SO4 (mg/L) 3 3 1.27 3.84 2.03

Sr (μg/L) 3 3 16.7 18.4 17.5

10 Helotes Creek Ca (mg/L) 7 7 12.0 90.0 40.8

Mg (mg/L) 7 7 0.92 16.4 7.84

Na (mg/L) 7 7 0.46 23.4 8.67

HCO3 (mg/L) 2 2 122 248 ND

Cl (mg/L) 7 7 0.64 41.5 16.7

SO4 (mg/L) 7 7 1.64 56.7 24.1

Sr (μg/L) 6 6 54.3 275 136
Geochemical Conditions 31

Groundwater

Major-ion concentrations measured in groundwater 

samples varied both temporally and spatially (figs. 8A–G). 

Concentrations in samples from the urban and unconfined 

wells (Parkwood and Shavano) were generally the most 

variable, which was evident during the wet period. This 

variability is associated with recharge during the wet 

period. In contrast, major-ion concentrations in samples 

collected from the rural and unconfined site (Seco well) 

were generally less variable. This contrast in the variability 

of major-ion geochemistry among the unconfined sites is 

consistent with the contrast in the variability in SC among 

these sites (figs.5A–C) also indicating that Seco well is less 

connected to localized groundwater recharge sources than 

are Parkwood and Shavano wells. Major-ion concentrations 

measured in the confined Mission well were distinct from 

those at the other wells, with much higher concentrations 

of Cl and, in particular, Sr (figs. 8E, G). These differences 

in concentration of major ions are consistent with mixing 

with a small component of saline groundwater at the 

Mission well. 

Changes in major-ion geochemistry over the multiyear 

study period were evident at only a few sites and for only a 

few compounds (figs. 8A–G). Mg and SO

4

 concentrations 

increased at the unconfined Shavano well, although a 

similar change was not observed at Parkwood well. The 

changes in Mg and SO

4

 at Shavano well were associated 

with changes in environmental concentrations rather than 

sampling variability given the very low range in RPDs 

associated with sample processing (table 4). Major-ion 

concentrations at Zarzamora and Turtle wells, which are 

both relatively deep confined wells, were generally similar 

to each other and relatively constant, with no notable 

temporal variability. At Comal 1 spring site, concentrations 

of sodium (Na), Cl, and SO

4

 decreased slightly over the 

study period. Although the magnitude of change in Na and 

Cl was smaller than that for SO

4

, only part of the variability 

in Na and Cl may be attributed to sample processing based 

on low RPDs observed for replicate samples (table 4), 

indicating that environmental changes, although variable, 

were occurring with these constituents. These gradual 

shifts to lower concentrations might reflect slow changes 

in regional flow paths in response to increased recharge 

containing a higher proportion of younger water. Minor 

temporal changes in major-ion concentrations at the 

confined groundwater sites (Zarzamora well, Turtle well, 

and Comal 1 spring) contrast with the enhanced variability 

at shallow recharge-zone sites (Parkwood well and Shavano 

well) that was evident during periods of rapid groundwater 

recharge, highlighting the importance of periods of rapid 

hydrologic change in contributing groundwater chemistry.
32 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

200

250

300

350

400

6

10

14

18

Bicarbonate, in milligrams per liter Sodium, in milligrams per liter Magnesium, in milligrams per liter Calcium, in milligrams per liter

5

10

15

20

60

70

80

90

100

110

120

130

A

B

C

D

Dry period1 Wet period2

Groundwater sample, 

 by site (fig. 1; table 1)

 Seco well

 Parkwood well

 Shavano well

 Zarzamora well

 Turtle well

 Mission well

 Comal 1 spring

EXPLANATION

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the “wet period.”

Figure 8. Concentrations of A, calcium, B, magnesium, C, sodium, D, bicarbonate, E, chloride, F, sulfate, and G, strontium measured in 

groundwater samples and Comal 1 spring samples collected from the San Antonio segment of the Edwards aquifer, south-central Texas, 

July 2013–December 2016.
Geochemical Conditions 33

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Strontium, in micrograms per liter Sulfate, in milligrams per liter Chloride, in milligrams per liter

5

10

15

20

25

30

0

10

20

30

40

50

0

400

800

1,200

1,600

2,000

E

F

G

Dry period1 Wet period2

Groundwater sample,

 by site (fig. 1; table 1)

 Seco well

 Parkwood well

 Shavano well

 Zarzamora well

 Turtle well

 Mission well

 Comal 1 spring

EXPLANATION

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the “wet period.”

Figure 8. Concentrations of A, calcium, B, magnesium, C, sodium, D, bicarbonate, E, chloride, F, sulfate, and G, strontium measured in 

groundwater samples and Comal 1 spring samples collected from the San Antonio segment of the Edwards aquifer, south-central Texas, 

July 2013–December 2016.—Continued
34 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Nutrient Geochemistry

Rainfall

Rainfall samples often contained measurable 

concentrations of NO3

, NO2, DON, and ortho-P at both rural 

and urban rainfall sites, indicating that atmospheric deposition 

is a source of N and P to the study area (table 7). Median 

concentrations of total-N were 0.30 mg/L (range = 0.13–

0.65 mg/L) and 0.49 mg/L (range = <0.10–0.59 mg/L) for 

rural (n = 4) and urban (n = 9) rainfall samples, respectively. 

At both urban and rural rainfall sites, NH4

 consistently had 

the highest median concentration among N species, followed 

by NO3

 and DON. Concentrations of ortho-P were low or not 

detectable in many samples, and the highest concentration 

(0.023 mg/L) was measured in a sample of urban rainfall. 

These results are consistent with the range in ortho-P 

concentrations (<0.004–0.077 mg/L) measured in rainfall 

samples collected in south-central Texas in 2015–16 (Lambert 

and others, 2017). A broad range in ortho-P concentrations in 

regional rainfall samples indicates that, at times, atmospheric 

deposition of ortho-P should be considered a potential source 

input to the aquifer.

Surface Water

Relative proportions of N species in surface water (NH4, 

NO3

, and DON) varied among surface-water sites. NO3 was 

detected in all surface-water samples, ranging in concentration 

from 0.05 to 2.02 mg/L, and was the dominant N species 

(based on median values) at most surface-water sites (table 8). 

The exception was Culebra Creek, where the dominant N 

species was DON. DON was the second highest concentration 

of N species at Helotes Creek; a median value could not be 

calculated for Seco Creek. NH4

 concentrations at all surfacewater sites were usually low relative to NO3

 and DON, but 

ranged from <0.01 to 0.06 mg/L. 

Table 7. Summary of selected nutrients and stable nitrogen and oxygen isotopes of nitrate (delta nitrogen-15 in nitrate and delta 

oxygen-18 in nitrate) measured in rainfall samples collected from Frio River (representative of rural land cover) and Urban rainfall 1, 2, 

and 3, San Antonio segment of the Edwards aquifer, south-central Texas, 2012–16.

[LRL, laboratory reporting level; NH4-N, ammonia as nitrogen; mg/L, milligram per liter; NO3-N, nitrate as nitrogen; NO2-N, nitrite as nitrogen; <, less than; 

ND, not determined; DON, dissolved organic nitrogen; ortho-P, orthophosphate as phosphorus; total-N, total nitrogen; δ15N-NO3, delta nitrogen-15 in nitrate; 

δ18O-NO3, delta oxygen-18 in nitrate]

Map 

identifier 

(fig. 1; 

table 1)

Short name 

(fig. 1; table 1) Constituent Number of samples

Number of concentrations equal 

to or greater than 

the LRL

Minimum 

value

Maximum 

value

Median 

value

7 Frio River1 NH4

-N (mg/L) 4 4 0.07 0.41 0.18

NO3

-N (mg/L) 4 4 0.04 0.16 0.11

NO2

-N (mg/L) 4 1 <0.001 0.002 ND

DON (mg/L) 4 4 0.01 0.08 0.02

Ortho-P (mg/L) 4 3 <0.004 0.005 ND

Total-N (mg/L) 4 4 0.13 0.65 0.30

δ15N-NO3

 (per mil) 3 3 −4.40 0.18 −3.91

δ18O-NO3

 (per mil) 3 3 64.42 68.65 64.48

4, 5, 6 Urban rainfall 

1, 2, 3

NH4

-N (mg/L) 9 9 0.04 0.73 0.25

NO3

-N (mg/L) 9 9 0.03 0.47 0.12

NO2

-N (mg/L) 9 6 <0.001 0.003 0.003

DON (mg/L) 4 3 <0.03 0.18 0.05

Ortho-P (mg/L) 9 5 <0.004 0.023 0.004

Total-N (mg/L) 4 3 <0.10 0.59 0.49

δ15N-NO3

 (per mil) 8 8 −5.28 1.94 −1.33

δ18O-NO3

 (per mil) 8 8 54.20 69.05 62.03

1

Rainfall samples collected at U.S. Geological Survey streamflow-gaging station 08195000 Frio River at Concan, Texas (Frio River site), were considered 

representative of rural rainfall.
Nutrient Geochemistry 35

Table 8. Summary of selected nutrients and stable nitrogen and oxygen isotopes of nitrate (delta nitrogen-15 in nitrate and delta 

oxygen-18 in nitrate) measured in surface-water samples collected from Frio River, Seco Creek, Culebra Creek, and Helotes Creek, San 

Antonio segment of the Edwards aquifer, south-central Texas, 2010–16.

[LRL, laboratory reporting level; NH4-N, ammonia as nitrogen; mg/L, milligram per liter; <, less than; ND, not determined; NO3-N, nitrate as nitrogen; NO2-N, 

nitrite as nitrogen, DON, dissolved organic nitrogen; ortho-P, orthophosphate as phosphorus; total-N, total nitrogen; δ15N-NO3, delta nitrogen-15 in nitrate; 

δ18O-NO3, delta oxygen-18 in nitrate]

Map 

identifier 

(fig. 1; 

table 1)

Short name 

(fig. 1; table 1) Constituent Number of samples

Number of 

concentrations 

equal to or 

greater than 

the LRL

Minimum 

value

Maximum 

value

Median 

value

7 Frio River NH4-N (mg/L) 46 16 <0.01 0.06 ND

NO3-N (mg/L) 46 46 0.05 1.68 0.40

NO2-N (mg/L) 46 28 <0.001 0.002 ND

DON (mg/L) 36 12 <0.03 0.45 ND

Ortho-P (mg/L) 46 9 <0.004 0.006 ND

Total-N (mg/L) 20 20 0.17 1.80 0.69

δ15N-NO3 (per mil) 26 26 1.56 8.60 6.30

δ18O-NO3 (per mil) 26 26 0.29 14.1 4.73

8 Seco Creek NH4-N (mg/L) 2 2 <0.01 0.02 0.02

NO3-N (mg/L) 2 2 0.16 0.35 0.25

NO2-N (mg/L) 2 2 0.004 0.008 0.006

DON (mg/L) 2 1 ND1 0.45 ND

Ortho-P (mg/L) 2 2 0.024 0.036 0.030

Total-N (mg/L) 2 2 0.63 0.95 0.79

δ15N-NO3 (per mil) 2 2 1.19 3.13 2.16

δ18O-NO3 (per mil) 2 2 11.5 21.6 16.3

9 Culebra Creek NH4-N (mg/L) 3 3 0.01 0.03 0.02

NO3-N (mg/L) 3 3 0.26 2.02 0.42

NO2-N (mg/L) 3 3 0.004 0.014 0.004

DON (mg/L) 3 3 0.55 0.87 0.61

Ortho-P (mg/L) 3 3 0.033 0.061 0.034

Total-N (mg/L) 3 3 0.85 2.90 1.00

δ15N-NO3 (per mil) 2 2 0.01 0.38 0.20

δ18O-NO3 (per mil) 2 2 6.65 7.06 6.86

10 Helotes Creek NH4-N (mg/L) 8 6 <0.01 0.05 0.03

NO3-N (mg/L) 8 8 0.30 1.54 0.56

NO2-N (mg/L) 8 8 0.007 0.020 0.010

DON (mg/L) 7 5 <0.27 0.51 0.27

Ortho-P (mg/L) 8 8 0.007 0.105 0.016

Total-N (mg/L) 6 6 0.58 2.10 0.82

δ15N-NO3 (per mil) 7 7 1.14 6.62 4.20

δ18O-NO3 (per mil) 7 7 6.49 13.7 7.79

1

A concentration of <0.59 was measured in one of the two samples when the LRL was 0.59 mg/L.
36 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Seco, Culebra, and Helotes Creeks are ephemeral and 

flow only in response to rainfall events; therefore, fewer 

samples were collected at these sites than at the Frio River 

(table 8). The lowest median daily discharge at the Frio 

River during this study was 5.49 ft3/s on September 7, 2013 

(fig. 4). Samples were collected at the Frio River during both 

base-flow and stormwater-runoff conditions, although more 

samples were collected during base-flow conditions. As a 

result, nutrient species concentrations varied over a wide 

range of flow conditions in comparison to other surfacewater sites. NO3

 concentrations measured in samples from 

the Frio River ranged from 0.05 to 1.68 mg/L (table 8), 

although slightly higher values (about 2.0 mg/L) were 

measured by the continuous NO3 sensor (discussed in the 

section “Continuous Nitrate Monitoring”). The median NO3

concentrations at surface-water sites (ranging from 0.25 

to 0.56 mg/L; table 8) were similar to or higher than the 

estimated national background concentration for streams 

of 0.24 mg/L (Dubrovsky and others, 2010). The highest 

median NO3 concentration was at Helotes Creek (0.56 

mg/L; table 8), which is about twice the estimated national 

background concentration for streams in undeveloped areas 

(0.24 mg/L) but less than the estimated national background 

concentration for urban streams (about 0.8 mg/L) (Dubrovsky 

and others, 2010). This higher concentration at Helotes 

Creek might result from a contribution of NO3 from urban 

sources. The estimated national background for streams in 

agricultural areas is about 2.7 mg/L, and the excess NO3 in 

these streams is attributed primarily to fertilizer application 

(Dubrovsky and others, 2010). Median NO3 concentrations 

for both rural (0.25–0.42 mg/L) and urban (0.56 mg/L) 

surface-water sites examined in this study (table 8) were 

substantially lower than the national background concentration 

for streams in agricultural areas (about 2.7 mg/L; Dubrovsky 

and others, 2010) and are therefore consistent with minor 

inputs of agricultural fertilizer. 

Ortho-P was detected in all samples collected from 

Seco, Culebra, and Helotes Creeks, but in only about 

20 percent of the samples from the Frio River (table 8). 

Median concentrations of ortho-P for Seco, Culebra, 

and Helotes Creeks ranged from 0.016 to 0.034 mg/L 

(table 8), slightly higher than the national background 

concentration of 0.010 mg/L (Dubrovsky and others, 2010). 

The median concentration of ortho-P for the Frio River was 

not determined; however, samples with detections were 

generally associated with higher flow conditions and likely 

had a storm-runoff component. These results indicate that 

the base-flow component of Frio River discharge, which is 

supported primarily by groundwater springs, typically has 

little to no ortho-P.

Groundwater

NO3 concentrations in groundwater samples were 

above the LRL, and NO3 was the most abundant N species 

at all groundwater sites (table 9). The lowest and highest 

median NO3 concentrations were measured at the rural 

Seco well (1.04 mg/L) and the urban Parkwood well 

(2.65 mg/L), respectively. In contrast, NH4 and DON were 

infrequently detected in groundwater, especially at the 

confined groundwater sites and at Comal 1 spring, where NH4

was detected only once (at Turtle well) and DON was not 

detected. Ortho-P was detected in all groundwater samples, 

with median concentrations ranging from 0.007 mg/L at Seco 

and Mission wells to 0.018 mg/L at Shavano well; ortho-P at 

Shavano well was substantially higher in concentration than 

at the other groundwater sites. Median ortho-P values at all 

of the groundwater sites were less than the estimated national 

background value for groundwater of 0.030 mg/L (Dubrovsky 

and others, 2010).

NO3 concentrations in groundwater samples (fig. 9A) 

varied much less than in surface-water samples (table 8), 

except at the unconfined urban Parkwood well during the wet 

period. Relatively low NO3 concentrations at Parkwood well 

were associated with samples collected in response to recharge 

events. For many of these samples, NO3 concentrations were 

less than 1.0 mg/L; between these storm-response samples, 

NO3 concentrations at Parkwood well returned to an ambient 

concentration between about 2.5 and 3.0 mg/L (fig. 9A). At 

Shavano well, few storm-response samples were collected, 

so most samples represent ambient groundwater conditions. 

NO3 concentrations at Shavano well were typically higher 

during the dry period (around 2 mg/L) and consistently lower 

during the wet period (around 1.6 mg/L). This decrease in 

NO3 concentration during the wet period likely reflects mixing 

with a larger component of recent, lower NO3 recharge water. 

Similar to Parkwood well, this site is characterized by a rapid 

response to rainfall and recharge events (fig. 5C). 

Temporal variability in ortho-P was generally small, but 

was apparent at several sites (fig. 9B). The most dynamic 

variability was at Shavano well, where ortho-P concentrations 

decreased from about 0.020 to 0.012 mg/L over the course of 

the study. Turtle and Mission wells also had small decreases in 

ortho-P concentrations, consistent with lower concentrations 

during the wet period. Longer term data that capture additional 

hydrologic variability would aid in determining if there are 

consistent temporal patterns in ambient ortho-P concentrations 

at these sites.
Nutrient Geochemistry 37

Table 9. Summary of selected nutrients and stable nitrogen and oxygen isotopes of nitrate (delta nitrogen-15 in nitrate and delta 

oxygen-18 in nitrate) measured in samples collected from groundwater wells and Comal 1 spring, San Antonio segment of the Edwards 

aquifer, south-central Texas, 2013–16.—Continued

[LRL, laboratory reporting level; NH4-N, ammonia as nitrogen; mg/L, milligram per liter; <, less than; ND, not determined; NO3-N, nitrate as nitrogen; NO2-N, 

nitrite as nitrogen, DON, dissolved organic nitrogen; ortho-P, orthophosphate as phosphorus; total-N, total nitrogen; δ15N-NO3, delta nitrogen-15 in nitrate; 

δ18O-NO3, delta oxygen-18 in nitrate]

Map 

identifier 

(fig. 1; 

table 1)

Short name

(fig. 1; table 1) Constituent Number of samples

Number of 

concentrations 

equal to or 

greater than 

the LRL

Minimum 

value

Maximum 

value

Median 

value

11 Seco well NH4-N (mg/L) 29 4 <0.01 0.23 ND

NO3-N (mg/L) 29 29 0.89 1.20 1.04

NO2-N (mg/L) 29 5 <0.003 0.016 ND

DON (mg/L) 16 1 <0.07 0.08 ND

Ortho-P (mg/L) 29 29 0.005 0.014 0.007

Total-N (mg/L) 16 1 <1.0 1.3 ND

δ15N-NO3 (per mil) 29 29 4.33 5.75 5.05

δ18O-NO3 (per mil) 29 29 3.54 4.91 3.93

12 Parkwood well NH4-N (mg/L) 21 1 <0.01 0.01 ND

NO3-N (mg/L) 32 32 0.79 3.12 2.65

NO2-N (mg/L) 32 4 <0.003 0.012 ND

DON (mg/L) 20 0 <0.07 ND ND

Ortho-P (mg/L) 32 32 0.007 0.013 0.010

Total-N (mg/L) 20 3 <1.2 2.1 1.200

δ15N-NO3 (per mil) 32 32 5.29 9.43 7.69

δ18O-NO3 (per mil) 32 32 4.76 6.14 5.09

13 Shavano well NH4-N (mg/L) 15 0 <0.01 ND ND

NO3-N (mg/L) 20 20 1.51 2.11 1.74

NO2-N (mg/L) 20 20 <0.001 0.002 ND

DON (mg/L) 14 0 <0.005 ND ND

Ortho-P (mg/L) 15 15 0.012 0.020 0.018

Total-N (mg/L) 0 0 ND ND ND

δ15N-NO3 (per mil) 10 10 5.77 6.25 6.02

δ18O-NO3 (per mil) 10 10 5.70 6.49 6.20

14 Zarzamora well NH4-N (mg/L) 27 0 <0.01 ND ND

NO3-N (mg/L) 27 27 1.78 1.94 1.83

NO2-N (mg/L) 27 3 <0.003 0.012 ND

DON (mg/L) 16 0 <0.07 ND ND

Ortho-P (mg/L) 27 27 0.007 0.011 0.008

Total-N (mg/L) 16 0 ND ND ND

δ15N-NO3 (per mil) 26 26 7.82 8.40 8.13

δ18O-NO3 (per mil) 26 26 4.76 5.96 5.14

Table 9. Summary of selected nutrients and stable nitrogen and oxygen isotopes of nitrate (delta nitrogen-15 in nitrate and delta 

oxygen-18 in nitrate) measured in samples collected from groundwater wells and Comal 1 spring, San Antonio segment of the Edwards 

aquifer, south-central Texas, 2013–16.

[LRL, laboratory reporting level; NH4-N, ammonia as nitrogen; mg/L, milligram per liter; <, less than; ND, not determined; NO3-N, nitrate as nitrogen; NO2-N, 

nitrite as nitrogen, DON, dissolved organic nitrogen; ortho-P, orthophosphate as phosphorus; total-N, total nitrogen; δ15N-NO3, delta nitrogen-15 in nitrate; 

δ18O-NO3, delta oxygen-18 in nitrate]
38 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Table 9. Summary of selected nutrients and stable nitrogen and oxygen isotopes of nitrate (delta nitrogen-15 in nitrate and delta 

oxygen-18 in nitrate) measured in samples collected from groundwater wells and Comal 1 spring, San Antonio segment of the Edwards 

aquifer, south-central Texas, 2013–16.—Continued

[LRL, laboratory reporting level; NH4-N, ammonia as nitrogen; mg/L, milligram per liter; <, less than; ND, not determined; NO3-N, nitrate as nitrogen; NO2-N, 

nitrite as nitrogen, DON, dissolved organic nitrogen; ortho-P, orthophosphate as phosphorus; total-N, total nitrogen; δ15N-NO3, delta nitrogen-15 in nitrate; 

δ18O-NO3, delta oxygen-18 in nitrate]

Map 

identifier 

(fig. 1; 

table 1)

Short name

(fig. 1; table 1) Constituent Number of samples

Number of 

concentrations 

equal to or 

greater than 

the LRL

Minimum 

value

Maximum 

value

Median 

value

15 Turtle well NH4-N (mg/L) 15 1 <0.01 0.01 ND

NO3-N (mg/L) 20 20 1.98 2.27 2.09

NO2-N (mg/L) 20 1 <0.001 0.001 ND

DON (mg/L) 13 0 <0.01 ND ND

Ortho-P (mg/L) 15 15 0.009 0.014 0.010

Total-N (mg/L) 0 0 ND ND ND

δ15N-NO3 (per mil) 12 12 7.33 7.70 7.57

δ18O-NO3 (per mil) 12 12 5.18 5.78 5.40

16 Mission well NH4-N (mg/L) 15 0 <0.013 ND ND

NO3-N (mg/L) 20 20 1.25 1.67 1.48

NO2-N (mg/L) 20 2 <0.001 0.003 ND

DON (mg/L) 12 0 <0.02 ND ND

Ortho-P (mg/L) 15 15 0.005 0.010 0.007

Total-N (mg/L) 0 0 ND ND ND

δ15N-NO3 (per mil) 12 12 5.64 5.96 5.79

δ18O-NO3 (per mil) 12 12 3.69 4.57 3.97

18 Comal 1 spring NH4-N (mg/L) 27 0 <0.01 0.04 ND

NO3-N (mg/L) 27 27 1.82 2.16 1.92

NO2-N (mg/L) 27 2 <0.003 0.013 ND

DON (mg/L) 16 0 <0.07 ND ND

Ortho-P (mg/L) 27 27 0.007 0.012 0.009

Total-N (mg/L) 16 1 <2.0 2.0 ND

δ15N-NO3 (per mil) 27 27 6.68 7.84 6.96

δ18O-NO3 (per mil) 27 27 4.79 5.83 5.22
Nutrient Geochemistry 39

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Orthophosphate-phosphorous,

in milligrams per liter

Nitrate-nitrogen,

in milligrams per liter

0

0.005

0.010

0.015

0.020

0.025

0.5

1.0

1.5

2.0

2.5

3.0

3.5

A

B

Dry period1 Wet period2

Groundwater sample,

 by site (fig. 1; table 1)

 Seco well

 Parkwood well

 Shavano well

 Zarzamora well

 Turtle well

 Mission well

 Comal 1 spring

EXPLANATION

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the “wet period.”

Figure 9. Concentrations of A, nitrate-nitrogen and B, orthophosphate-phosphorous measured in water samples collected from 

groundwater wells and Comal 1 spring, San Antonio segment of the Edwards aquifer, south-central Texas, July 2013–December 2016.
40 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Continuous Nitrate Monitoring

Surface Water

Continuous NO3 monitoring, which began at the Frio 

River in September 2015 (fig. 10), provides more detail about 

temporal NO3

 variability than do the discrete samples (fig. 9). 

Comparison of NO3

 concentrations and discharge showed a 

consistent pattern of low NO3

 concentrations measured during 

base-flow conditions and higher NO3

 concentrations measured 

during high-flow conditions. A large rainfall-recharge event 

in September 2016 resulted in the highest discharge at the 

Frio River during the period of continuous monitoring and 

illustrates the consistent relations between NO3

 concentration, 

streamflow, and SC (fig. 11). As discharge increased initially, 

SC decreased rapidly, consistent with rapid runoff of lowconductance stormwater-runoff from rainfall. During the 

initial rise in discharge, an increase in turbidity was evident 

based on visual observations, and a corresponding gap in the 

continuous NO3

 data occurred. NO3 sensor measurements 

resumed about midway through the discharge peak, recording 

a corresponding steady and rapid increase in concentration 

from about 0.4 to 2.0 mg/L. The NO3

 increase was temporally 

delayed relative to the discharge increase, occurring when the 

storm pulse was receding. This lag in the NO3

 concentration 

increase is hypothesized to primarily result from a delay in 

NO3

 mobilization from soils. Prior studies in other watersheds 

(Lucey and Goolsby, 1993; Reynolds and Edwards, 1995; 

Rusjan and others, 2008) have recorded NO3

 increases during 

Date

Sept. 2015

Jan. 2016

May 2016

Sept. 2016

Jan. 2017

5

50

500

5,000

Daily median discharge, in cubic feet per second

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6

1.8

2.0

Daily median nitrate-nitrogen, in milligrams per liter

Frio River (fig. 1; table 1)

Daily median nitrate-nitrogen—Gray shading

indicates 95 percent confidence interval

Daily median discharge

EXPLANATION

Figure 10. Daily median discharge and daily median nitrate-nitrogen concentration at Frio River (station 0819500), San Antonio 

segment of the Edwards aquifer, south-central Texas, September 2015–December 2016. The shaded gray area associated with nitratenitrogen concentrations represents the probability range in error (plus or minus 0.07 milligrams per liter) with a 95 percent confidence 

interval. Gaps in the continuous record indicate missing data.
Nutrient Geochemistry 41

Date Oct. 15, 2016

Sept. 20, 2016

Sept. 25, 2016

Sept. 30, 2016

Oct. 5, 2016

Oct. 10, 2016

Oct. 20, 2016

100

1,000

10,000

Instantaneous discharge, in cubic feet per second

Nitrate-nitrogen, in milligrams per liter

0.2

0.6

1.0

1.4

1.8

2.2

Specific conductance, in microsiemens

per centimeter at 25 degrees Celsius

625

550

475

400

250

325

Frio River (fig. 1; table 1)

Nitrate-nitrogen—Gray shading indicates

95 percent confidence interval

Instantaneous discharge

Specific conductance

EXPLANATION

Figure 11. Instantaneous discharge, nitrate-nitrogen concentration, and specific conductance for a single stormwater-runoff event 

at Frio River (station 0819500), San Antonio segment of the Edwards aquifer, south-central Texas, September 20, 2016–October 20, 2016. 

The shaded gray area associated with nitrate-nitrogen concentrations represents the probability range in error (plus or minus 0.07 

milligrams per liter) with a 95 percent confidence interval. Gaps in the continuous record indicate missing data.

the recession that follows peak discharge, attributing this pattern 

to NO3 mobilization from soils. As the discharge peak at the 

Frio River receded, NO3 concentrations gradually declined to 

near, although somewhat higher than, base-flow values prior to 

the event (fig. 10).

Unconfined Rural Groundwater

Continuous NO3 monitoring at Seco well began in July 

2014, during the dry period when WLA was gradually declining 

(fig. 12A). Several small recharge events (less than 2 in. of 

rainfall) between July 2014 and March 2015 resulted in small 

(less than 3 ft) increases in WLA and corresponding small but 

rapid increases and decreases in NO3 concentration. Following 

the onset of more frequent and larger rain events during the wet 

period—including multiple events in spring 2015, winter 2015, 

and spring 2016—NO3 concentration increased a small amount 

(approximately 0.15 mg/L) and then remained at that higher 

concentration. The total variability in NO3 at Seco well was 

relatively small, ranging from about 1.0 to 1.3 mg/L. Although 

the difference in NO3 concentration between the dry period and 

the wet period was small, it is distinguishable outside of the 

NO3 sensor’s precision of plus or minus 0.07 mg/L (Opsahl and 

others, 2017), indicating that NO3 concentrations at this site 

responded to longer term changes in hydrologic conditions. 

Unconfined Urban Groundwater

As at Seco well, NO3 continuous monitoring at Parkwood 

well began during the dry period (September 2014), and 

the transition from the dry to wet period showed varying 

responses to changing hydrologic conditions (fig. 12B). 

Ambient NO3 concentrations at Parkwood well were initially 

high (about 2.7 mg/L) in comparison to Seco well (about 

1.0 mg/L), as was the overall variability in NO3 concentration 

at Parkwood well (about 0.8 to 3.3 mg/L) in comparison 

to Seco well (about 1.0 to 1.3 mg/L) (fig. 12A, B). NO3

concentrations at Parkwood well were relatively constant 

during the dry period, except for a small event in November 

2014—several inches of rain and corresponding recharge 

resulted in a small increase in WLA and a corresponding 

slight decrease in NO3 concentration. In response to frequent 

rainfall during the transition to the wet period in spring 2015, 

WLA rose rapidly by about 40 ft over a period of a few days, 

and concurrently, NO3 concentration decreased from about 

2.7 mg/L to less than 1 mg/L. A similar pattern of rapid WLA 

increase and corresponding NO3 concentration decrease 

occurred during several subsequent rainfall and recharge 

events during the remainder of 2015 and in 2016. After each 

of these events, as WLA stabilized at a new and higher value, 

NO3 concentrations typically returned to a consistent ambient 

concentration of around 2.7 mg/L. These rapid rises in WLA 
42 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Date

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

0

1

2

3

4

Daily rainfall,

in inches Daily median nitrate-nitrogen, in milligrams per liter

Daily median water-level altitude, in feet above

North American Vertical Datum of 1988

760

776

792

808

824

840

0.9

1.0

1.1

1.2

1.3

1.4

A

Seco well (fig. 1; table 1)

Rainfall data 2 site (fig. 1; table 1)

Dry period1 Wet period2

Dry period1 Wet period2

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the ”wet period.”

Daily median nitrate-nitrogen—Gray shading

indicates 95 percent confidence interval

Daily median water-level altitude

EXPLANATION

Figure 12. Rainfall, daily median water-level altitude, and daily median nitrate-nitrogen for A, Seco well, July 2014–December 2016, 

and B, Parkwood well, September 2014–December 2016, San Antonio segment of the Edwards aquifer, south-central Texas. The shaded 

gray area associated with nitrate-nitrogen concentrations represents the probability range in error (plus or minus 0.07 milligrams per 

liter) with a 95 percent confidence interval. Gaps in the continuous record indicate missing data.
Nutrient Geochemistry 43

Date

Sept. 2014

Mar. 2015

Sept. 2015

Mar. 2016

Sept. 2016

Dec. 2016

3.4

3.2

3.0

2.8

2.6

2.4

2.2

2.0

1.8

1.6

1.4

1.2

1.0

0.8

0.6 670

690

710

730

680

700

720

740

750

Daily median nitrate-nitrogen, in milligrams per liter

Daily median water-level altitude, in feet above

North American Vertical Datum of 1988

0

1

2

3

4

Daily rainfall,

in inches 

B

Rainfall data 1 site (fig. 1; table 1)

Parkwood well (fig. 1; table 1)

Dry period1 Wet period2

Dry period1 Wet period2

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the ”wet period.”

Daily median nitrate-nitrogen—Gray shading

indicates 95 percent confidence interval

Daily median water-level altitude

EXPLANATION

Figure 12. Rainfall, daily median water-level altitude, and daily median nitrate-nitrogen for A, Seco well, July 2014–December 2016, and 

B, Parkwood well, September 2014–December 2016, San Antonio segment of the Edwards aquifer, south-central Texas. The shaded gray 

area associated with nitrate-nitrogen concentrations represents the probability range in error (plus or minus 0.07 milligrams per liter) 

with a 95 percent confidence interval. Gaps in the continuous record indicate missing data.—Continued
44 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

and sharp decreases in NO3 concentration are hypothesized 

to result from mixing of groundwater with a rapid influx of 

more dilute, low-NO3 recharge water. NO3 concentrations at 

Helotes Creek, which might represent a suitable surrogate 

for urban streams recharging in the area, ranged from 0.30 to 

1.54 mg/L (table 8). Mixing between ambient groundwater 

and lower NO3 recharging stormwater runoff derived directly 

from recharge streams such as Helotes Creek can account 

for the rapid and short-term dilution of ambient water at 

Parkwood well.

Distinguishing Nitrate Sources

Measuring the isotopes δ15N-NO3 and δ18O-NO3 found 

in rainfall, surface-water, and groundwater samples provided 

insights into the sources contributing NO3 to the Edwards 

aquifer. Rainfall samples had characteristically low δ15N-NO3

and high δ18O-NO3 values (fig. 13A), which were within 

the range of previously published values for rainfall in this 

region (Lambert and others, 2017). Surface-water samples 

had a large range of δ15N-NO3 and δ18O-NO3 values, likely 

reflecting varying contributions of NO3 from different 

possible sources, including rainfall, inorganic fertilizers, 

human and animal waste, and natural soil NO3 (fig. 13A). 

Most of the δ15N-NO3 values measured in surface-water 

samples were less than 8 per mil, which indicates that little 

to no NO3 was likely derived from human or animal waste, 

which is characterized by relatively high δ15N-NO3 values 

(fig. 13A). Five surface-water samples had δ15N-NO3 values 

less than 5 per mil and δ18O-NO3 values greater than 10 

per mil (fig. 13B), which is consistent with mixing with a 

component of runoff from recent rainfall. Contributions to 

surface water from other low δ15N-NO3 sources, however, 

such as inorganic fertilizer (fig. 13A), cannot be ruled out on 

the basis of only isotopes. Groundwater samples generally 

have a much smaller range of isotopic values at each site, 

indicating less temporal variability at groundwater sites 

relative to surface water and rainfall, and results for each 

groundwater site tend to cluster in a distinct grouping 

(fig. 13B). One exception was a subset of samples from 

Parkwood well that had the lowest δ15N-NO3 values for this 

site (less than 7 per mil) (fig. 13B). This subset of samples 

was collected in response to rainfall and recharge events 

and, based on SC changes (fig. 5B), reflect mixing with 

a component of recent recharge water; the low δ15N-NO3

values for these samples correspond with low NO3

concentrations described in the “Nutrient Geochemistry” 

section of this report.

Isotopic values of NO3 in combination with NO3

concentrations can provide insight into NO3 sources (Kendall 

and others, 2014; Musgrove and others, 2016) (fig. 14). 

Among the groundwater sites, the small range of δ15N-NO3

values for Seco well (4.33 to 5.75 per mil) and corresponding 

low NO3 concentration (median NO3 of 1.04 mg/L) (table 9) 

indicate predominantly natural, soil-derived NO3 with 

little contribution from anthropogenic sources. In contrast, 

Parkwood well (with the exception of samples collected in 

response to recharge events) had the highest δ15N-NO3 values 

(median of 7.69 per mil) and NO3 concentrations (median of 

2.65 mg/L) (table 9), which is consistent with mixing with 

a component of NO3 from human or animal waste sources. 

Other groundwater sites were intermediate in their NO3

isotopic and concentration values, with higher δ15N-NO3

values corresponding to higher NO3 concentrations. On the 

basis of the observed patterns among wells (fig. 14), higher 

δ15N-NO3

 values and NO3 concentrations indicate a larger 

component of NO3 sourced from human and animal waste. 

Samples from Zarzamora well had somewhat higher δ15N-NO3

values than the general trend, which also might indicate a 

proportionally larger local component of NO3 derived from 

human and animal waste in comparison to the other confined 

groundwater sites. 

Factors Affecting Nitrate 

An examination of the concentration patterns of 

individual N species helps to explain controls on NO3

concentrations throughout the Edwards aquifer system. 

Median values indicate that NO3 in rainfall is generally 

low (median values for both rural and urban rainfall around 

0.1 mg/L) (table 7), is higher in surface water (median values 

ranging from 0.25 to 0.56 mg/L) (table 8), and is highest in 

groundwater (median values ranging from 1.04 to 2.65 mg/L) 

(table 9). The much lower median NO3 concentrations in 

rainfall (table 7) relative to surface water (table 8) indicate 

that NO3 in rainfall is generally a variable but relatively 

minor direct contributor of NO3 to groundwater. In contrast, 

NH4 and DON in rainfall contribute to the input of total N to 

the aquifer. Concentrations of NH4 were generally higher in 

rainfall than in surface water, and NH4 was rarely detected in 

groundwater (fig. 15A). In contrast, DON concentrations were 

generally lower in rainfall and generally higher in surface 

water (fig. 15B). DON concentrations in surface water are 

likely attributable to mobilization of DON from soils into 

the streams during storm events (Campbell and others, 2000; 

Cooper and others, 2007). However, like NH4, DON was 

detected in only a few groundwater samples.
Nutrient Geochemistry 45

0 2 4 6 8 10

Delta nitrogen-15 of nitrate, in per mil

–15 –10 –5 0 5 10 15 20 25 30 35 40

Delta nitrogen-15 of nitrate, in per mil

0

4

8

12

16

20

Delta oxygen-18 of nitrate, in per mil

-20

0

20

40

60

80

100

Delta oxygen-18 of nitrate, in per mil

Ammonium in Human and animal waste

fertilizer and

rainfall

Soil nitrogen

Nitrate in fertilizer

Nitrate in rainfall

A

B

EXPLANATION

Sample, by site (fig. 1; table 1)

Frio River

Seco Creek

Culebra Creek

Helotes Creek

Seco well

Parkwood well

Shavano well

Zarzamora well

Turtle well

Mission well

Comal 1 spring

EXPLANATION

Sample, by site (fig. 1; table 1)

Rural rainfall—Sample collected at 

 Frio River site

Urban rainfall—Sample collected at 

 Urban rainfall 1, Urban rainfall 2, 

 and Urban rainfall 3

Frio River

Seco Creek

Culebra Creek

Helotes Creek

Seco well

Parkwood well

Shavano well

Zarzamora well

Turtle well

Mission well

Comal 1 spring

Figure 13. Stable nitrogen and oxygen isotopes of nitrate (delta nitrogen-15 of nitrate [δ15N-nitrate] and delta oxygen-18 of nitrate

[δ18O-nitrate]) measured in water samples collected from rainfall, surface-water, and groundwater sampling sites in the San Antonio 

segment of the Edwards aquifer, south-central Texas, 2011–16, for A, all samples within the full range of δ15N-nitrate and δ18O-nitrate 

values reported in this study and B, samples within a δ15N-nitrate range between 0 and 10 per mil and a δ18O-nitrate range between 0 

and 20 per mil. Boxes show the typical range of nitrate sources, modified from Kendall and others (2014). Ranges do not reflect specific 

south-central Texas sources but a compilation of measured values from numerous studies.
46 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

0 0.5 1.0 1.5 2.0 2.5 3.0 3.5

Nitrate-nitrogen, in milligrams per liter

Delta nitrogen-15

of nitrate, in per mil

3

4

5

6

7

8

9

10

Anthropogenic Natural

Natural Anthropogenic

EXPLANATION

Groundwater sample, by site

 (fig. 1; table 1)

Seco well

Parkwood well

Shavano well

Zarzamora well

Turtle well

Mission well

Comal 1 spring

Figure 14. Stable nitrogen isotopes of nitrate (delta nitrogen-15 of nitrate [δ15N-nitrate]) values versus nitrate-nitrogen concentration 

measured in groundwater water samples collected from sampling sites in the San Antonio segment of the Edwards aquifer, southcentral Texas, 2013–16. Arrows depict a progression of nitrate-nitrogen from minimally affected (natural) to more heavily influenced by 

anthropogenic sources.

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

Ammonia-nitrogen, in milligrams per liter 1

A constituent concentration is considered estimated by the laboratory when results are greater than the long-term method detection level (LT–MDL) and less than the LRL; that is, a detection is 

considered likely, but quantification is considered questionable. The remark code of “E” (estimated) is assigned by the laboratory for these values (Childress and others, 1999).

2

Rural rainfall sample collected at Frio River (fig. 1; table 1).

3

Urban rainfall sample collected at Urban rainfall 1, Urban rainfall 2, and Urban rainfall 3 (fig. 1; table 1).

EXPLANATION

Detection—Value equal to or

 greater than the laboratory

 reporting level (LRL) and

 estimated value1

Nondetection—Value less

 than the LRL

A

Urban rainfall3

Rural rainfall2

Culebra Creek

Comal 1 spring

Frio River

Helotes Creek

Seco Creek

Seco well

Parkwood well

Shavano well

Zarzamora well

Turtle well

Mission well

Site

Urban rainfall3

Rural rainfall2

Culebra Creek

Comal 1 spring

Frio River

Helotes Creek

Seco Creek

Seco well

Parkwood well

Shavano well

Zarzamora well

Turtle well

Mission well

Site

Dissolved organic nitrogen, in milligrams per liter

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

EXPLANATION

Detection—Value equal to or

 greater than the LRL and

 estimated value1

Nondetection—Value less

 than the LRL

B

Figure 15. Concentrations of A, ammonia-nitrogen and B, dissolved organic nitrogen measured in water samples collected from 

sampling sites in the San Antonio segment of the Edwards aquifer, south-central Texas, 2010–16. Samples with nondetections are plotted 

at the laboratory reporting level.
Nutrient Geochemistry 47

The absence or near absence of NH4 and DON in 

groundwater might result from their oxidation to NO3

 by 

nitrification, a microbial process common in soils and aquatic 

environments (Kendall and others, 2007). Nitrification might 

occur subaqueously if DON and NH4

 reach the saturated zone 

during periods of rapid recharge and oxygenated conditions 

are present (DeSimone and Howes, 1998), as occurs in karst 

aquifers. This process has been previously proposed for 

the Edwards aquifer, where nitrification of DON and NH4

can account for the increase in NO3

 observed in the aquifer 

relative to recharging surface water (Musgrove and others, 

2016). In the present study, the decrease in concentrations 

of DON and NH4

 from rainfall to surface-water recharge to 

groundwater is consistent with transformation by nitrification 

(fig. 15A). Although there are processes other than nitrification 

that transform NH4

 and DON, such as biological uptake, 

the relation between the sum of NH4

 and DON versus NO3

concentrations (fig. 16) indicates nitrification as a plausible 

process occurring in the aquifer. The δ15N-NO3

 data show 

a progression from rainfall (median δ15N-NO3

 = −3.91 to 

−1.33) (table 7) to surface water (median δ15N-NO3

 = 0.20 

to 6.30) (table 8) to groundwater (median δ15N-NO3

 = 5.05 

to 8.13) (table 9), which provides further support for the 

hypothesis that NH4

 and DON are transformed to NO3 as 

water recharges the aquifer. The increase in δ15N-NO3

values 

is consistent with progressively higher δ15N-NO3

 values that 

were measured during conversion of NH4

 to NO3 in soils 

(Hübner, 1986; Kendall, 1998). Furthermore, Musgrove and 

others (2016) reported a similar increase in δ15N-NO3 values 

during the transformation of NH4

 and DON to NO3 in the 

Edwards aquifer.

Nitrogen Inputs and Cycling in the

Edwards Aquifer

An analysis of NO3 concentrations, δ15N-NO3 values, 

and additional N species that affect NO3

 concentrations (that 

is, NH4

 and DON) is necessary for distinguishing between 

rural and urban N sources and estimating the magnitude 

of anthropogenic NO3

 source contributions to the Edwards 

aquifer (fig. 14). Results of this study indicate that there 

were different sources of NO3

 in rural and urban areas 

and that anthropogenic activities contribute to higher NO3

concentrations in groundwater in the urban San Antonio 

area. The differences between NO3

 inputs in rural and urban 

groundwater are illustrated by a comparison of continuous 

NO3 monitoring and δ15N-NO3

 results at the rural unconfined 

Seco well and the urban unconfined Parkwood well in 

response to recharge events. At the rural unconfined well, the 

median NO3 concentration was relatively low (1.04 mg/L)

(table 9), and changes in NO3 concentration in response to 

recharge events were small, indicating little difference in 

NO3 concentration between ambient groundwater and recent 

surface-water recharge (fig. 12A). Furthermore, the δ15N-NO3

values at the rural Seco well were consistent with natural 

soil NO3 sources (fig. 13). In contrast, at Parkwood well, the 

median NO3 concentration was more than twice as high (2.65 

mg/L) (table 9), and NO3 concentration notably decreased in 

response to surface-water recharge (fig. 12B). These results 

indicate that locally ambient groundwater NO3 is diluted 

by lower NO3 recharge water (fig. 12B). δ15N-NO3 values 

measured in ambient groundwater samples were consistent 

with a component of human and animal waste as a source 

of the NO3

, whereas δ15N-NO3 values measured in stormresponse samples were indicative of a more natural, soilderived NO3

 source within the urban setting. As the recharge 

pulse moved through the aquifer, NO3 concentrations 

returned to the higher ambient concentrations typical of 

this site. In the urban San Antonio area, possible sources of 

higher δ15N-NO3 values are treated wastewater, wastewater 

from septic systems, leaky sewer pipes, or other as yet 

unidentified waste sources (Dubrovsky and others, 2010).

Natural background NO3

 concentrations in groundwater 

nationally and within the Edwards aquifer have been 

previously estimated to be about 1 mg/L (Dubrovsky and 

others, 2010; Musgrove and others, 2016). The ambient 

NO3

 concentration of about 1 mg/L at the rural Seco well 

therefore is consistent with the hypothesis that the NO3 at

Seco well is predominantly from natural rather than 

anthropogenic sources. In contrast, NO3

 concentrations at 

Parkwood well (maximum NO3 of 3.12 mg/L) (table 9) 

indicate that about two-thirds of the NO3 in groundwater 

samples collected for this study might result from 

anthropogenic NO3 sources, assuming that 1 mg/L is the 

natural background concentration and the remainder (2.12 

mg/L) is anthropogenic. Although groundwater sites sampled 

for this study were selected to be representative of a range 

of water-quality conditions and potential anthropogenic 

sources, higher NO3 concentrations have been measured 

at other groundwater wells in the aquifer (Musgrove and 

others, 2010), and thus the proportion of anthropogenic 

NO3 might be higher at some locations.

Natural biogeochemical processes in the aquifer 

affect the fate and transport of all N species and play a 

role in determining the distribution of NO3

 concentrations 

throughout the aquifer. Loss of NO3

 through denitrification, 

a process that occurs under anoxic conditions, is unlikely 

to occur in the oxic conditions of the Edwards aquifer. 

However, as discussed previously, the process of nitrification 

of DON and NH4 from rainwater and surface water might be 

an important but previously unaccounted source of NO3 to 

Edwards aquifer groundwater, as evidenced by the decrease 

in DON plus NH4 concentrations from rainfall to surface 

water to groundwater and the corresponding increase in 

NO3 concentration (fig. 16). Characterization of sources, 

fate, and transport of NO3 in the San Antonio segment of 

the Edwards aquifer therefore requires consideration of 

all N species.
48 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

0 1.0 2.0 3.0 4.0

Nitrate-nitrogen, in milligrams per liter

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

Dissolved organic nitrogen plus ammonia nitrogen, in milligrams per liter

Laboratory reporting level

1

Detections consist of values greater than or equal to the laboratory reporting level (LRL) and estimated values. A constituent concentration is considered estimated by the laboratory

when results are greater than the long-term method detection level (LT–MDL) and less than the LRL; that is, a detection is considered likely, but quantification is considered questionable.

The remark code of “E” (estimated) is assigned by the laboratory for these values (Childress and others, 1999).

2

Nondetections are values reported for constituents that were analyzed for and were present at concentrations below the LRL and not quantified or were not present at all. For display

purposes only, nondetections are depicted on this graph at the concentration that defines the LRL.

Detection,1

 by site (fig. 1; table 1)

Rural rainfall—Sample collected at Frio

 River

Urban rainfall—Sample collected at 

 Urban rainfall 1, Urban rainfall 2, and

 Urban rainfall 3

Frio River

Seco Creek

Culebra Creek

Helotes Creek

Seco well

Parkwood well

Comal 1 spring

Nondetection2

 (ND), by site (fig. 1; 

 table 1)

Urban rainfall ND—Sample collected at

 Urban rainfall 1, Urban rainfall 2, and

 Urban rainfall 3

Seco well ND

Parkwood well ND

Zarzamora well ND

Comal 1 spring ND

EXPLANATION

Figure 16. The sum of dissolved organic nitrogen and ammonia-nitrogen concentration versus nitrate-nitrogen concentration 

measured in water samples collected from sampling sites in the San Antonio segment of the Edwards aquifer, south-central Texas, 

2010–16.

Pesticide Geochemistry

Rainfall

At least 1 pesticide was detected in each of the 

11 rainfall samples analyzed (table 10). A total of 

9 pesticides (6 herbicides and 3 insecticides) were detected. 

The herbicide dimethyl tetrachloroterephthalate (dacthal, 

abbreviated as DCPA) was detected the most frequently 

(8 of 11 samples; 73 percent), followed by the herbicides 

metolachlor and atrazine (7 of 11 samples; 64 percent). 

The DEA to atrazine ratio (DAR) is often used to evaluate 

the extent of atrazine degradation in the environment 

(Adams and Thurman, 1991; Thurman and Fallon, 1996; 

Townsend and Young, 2000). The DAR in rainfall was 

0.25–1.10 (median 0.53) for the 5 samples of rainfall 

in which both DEA and atrazine were detected (Opsahl 

and others, 2018), in the range typical for rainfall for the 

central United States (Scribner and others, 2005). Rainfall 

typically has a higher DAR than surface water because of 

the photodegradation of atrazine to DEA in the atmosphere 

(Thurman and Cromwell, 2000). Concentrations of pesticides 

in rainfall were low—the maximum concentration measured 

was 0.051 micrograms per liter (μg/L) (atrazine at Urban 

rainfall 2 site) (table 10).
Table 10. Summary of pesticide concentrations measured in rainfall samples collected from Frio River and Urban rainfall 1, 2, and 3 in the San Antonaquifer, south-central Texas, 2014–16. [Only those pesticides that were detected in at least one sample are shown. DEA, deethylatrazine; CIAT, 2-Chloro-4-isopropylamino-6-amino-s-triazine; μg/L, microgram per littetrachloroterephthalate; Hd, herbicide degradate; H, herbicide; I, insecticide; E, estimated; <, less than]

Map 

identifier 

(fig. 1; 

table 1)

Short name

(fig. 1; table 1) Date

DEA 

(CIAT) 

(μg/L)

Acetochlor 

(μg/L)

Atrazine 

(μg/L)

Carbaryl 

(μg/L)

Chlorpyrifos 

(μg/L)

DCPA 

(dacthal) 

(μg/L)

Malaoxon 

(μg/L)

Hd H H I I H I 

11 Frio River1 5/12/2015 E0.008 <0.010 0.015 <0.060 0.0095 <0.0076 <0.022 

5/13/2015 <0.010 <0.010 0.007 <0.060 0.0115 <0.0076 <0.022 

10/22/2015 <0.010 <0.010 <0.008 <0.060 <0.0100 0.0095 <0.022 

3/8/2016 <0.011 0.007 0.046 <0.060 <0.0100 0.0041 <0.022 

13 Urban rainfall 1 3/20/2015 <0.010 <0.010 <0.008 <0.060 <0.0100 0.0025 <0.022 

14 Urban rainfall 2 5/13/2015 E0.006 <0.010 0.017 E0.011 <0.0100 <0.0076 <0.022 

10/22/2015 <0.010 <0.010 <0.008 <0.060 <0.0100 0.0031 <0.022 

3/9/2016 E0.013 <0.010 0.051 <0.060 <0.0100 0.0032 <0.022 

15 Urban rainfall 3 5/9/2014 E0.011 <0.010 0.010 E0.007 <0.0116 0.0018 0.030 

5/12/2014 E0.007 <0.010 0.011 E0.004 <0.0100 0.0015 <0.022 

11/4/2014 <0.010 <0.010 <0.008 <0.060 <0.0100 0.0023 <0.022 

1Rainfall samples collected at U.S. Geological Survey streamflow-gaging station 08195000 Frio River at Concan, Texas (Frio River site), were considered representative of ru
50 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Surface Water

Sixteen pesticides (7 herbicides and 1 herbicide 

degradate; 5 insecticides and 3 insecticide degradates) each 

were detected in at least 1 surface-water sample (38 surfacewater samples were analyzed) (table 11). The herbicide 

metolachlor was detected most frequently (12 of

38 samples; 32 percent); concentrations of metolachlor 

were low (most were less than the LRL of 0.012 μg/L) and 

similar to concentrations measured in rainfall, with the 

exception of 1 urban rainfall sample that had a substantially 

higher concentration (fig. 17). For the 6 pesticides detected 

in both surface-water and rainfall samples, the range of 

concentrations measured in the 2 sample types was similar 

(Opsahl and others, 2018), and for 5 of the 6 pesticides 

(excepting carbaryl), the maximum concentration measured 

in rainfall exceeded that measured in surface water (fig. 18). 

However, 10 pesticides detected in surface water were 

not detected in rainfall (table 11). In surface water and 

groundwater, a lower DAR indicates atrazine that was more 

recently applied; a DAR greater than 1 indicates greater 

prevalence of the degradate than the parent compound, 

attributed to microbial breakdown of the parent compound 

in soil (Thurman and Fallon, 1996), and reflects little 

recent inputs of fresh (undegraded) atrazine. The DAR 

was 0.30–0.85 (median 0.75) for the 7 surface-water 

samples in which both DEA and atrazine were detected and 

were similar to the range in rainfall (0.25–1.10) (Opsahl 

and others, 2018), indicating that some of the surfacewater samples had a predominance of freshly applied 

atrazine but that other surface-water samples had a higher 

proportion of the degradate. 

Groundwater

Nine pesticides were detected in at least 1 of the 

115 groundwater samples collected (table 11). Among the 

unconfined wells, pesticides occurred much less frequently 

at the rural Seco well than at the 2 urban wells (Parkwood 

and Zarzamora wells). Three pesticides (atrazine, DEA, and 

simazine), all herbicides, each were detected in at least 1 

of 29 samples collected from the rural Seco well. Of these, 

DEA was detected in 55 percent of the samples (16 of 29); 

the other 2 pesticides were detected in just 1 (atrazine) 

or 2 (simazine) samples. All concentrations of DEA were 

estimated (less than the LRL of 0.010 μg/L). Most detections 

of DEA occurred during the first part of the study, in the 

dry period when WLA was low; DEA was not detected in 

any sample collected after May 2015, which was about 

2 weeks after WLA began to rise (fig. 19) and throughout the 

remainder of the wet period.

Nine pesticides each were detected in at least 1 of the 

59 total samples collected from the 2 urban wells (27 samples 

of confined groundwater at the deep Zarzamora well and 

32 samples of unconfined groundwater at the shallow 

Parkwood well) (table 11). Despite the difference in depths 

and hydrogeological character of the 2 urban wells, pesticide 

detection frequencies were similar, with the exception of 

prometon, which was detected only in samples collected at the 

shallow Parkwood well. The herbicides atrazine, DEA, and 

simazine were detected in more than 90 percent of samples 

collected at both wells (table 11). The herbicide prometon 

was detected in 61 percent (8 of 13) of the samples collected 

at Parkwood well during the dry period, but in only 7 percent 

(1 of 14) of the samples collected after May 2015 during 

the wet period (fig. 20A). In contrast to Parkwood well, the 

detection frequencies and concentrations at Zarzamora well 

changed little during contrasting dry and wet hydrologic 

conditions (fig. 20B). The herbicide metolachlor and the 

insecticides naphthol, azinphos-methyl, dieldrin, and fipronil 

sulfone (a degradate of fipronil) each were detected in just 

1 urban groundwater sample (table 11).

Pesticide detection frequency at Comal 1 spring was 

intermediate between that at the rural well and that at the 

two urban wells. Four herbicides or herbicide degradates—

atrazine, DEA, prometon, and simazine—were detected in 

at least 60 percent of the 27 samples collected at Comal 1 

spring (table 11), and DEA was detected in all but 1 sample 

(96 percent). No other pesticides were detected in the samples 

collected at Comal 1 spring. There were no temporal patterns 

in concentrations of atrazine, DEA, or simazine evident 

from the samples collected at Comal 1 spring (fig. 21). 

Concentrations of prometon, however, generally increased 

with the increase in spring discharge, which began in August 

2014, as hydrologic conditions transitioned between the dry 

and wet periods, but then decreased after July 2016, despite a 

continued increase in discharge.

The DAR was 0.10–3.0 (median = 1.86) for the 

115 groundwater samples in which both atrazine and DEA 

were detected (Opsahl and others, 2018). Two samples 

collected from Parkwood well yielded DARs that were 

substantially less than 1 (0.10 and 0.27), coinciding with 

relatively elevated concentrations of atrazine (0.26 and 

0.071 μg/L, respectively) (fig. 22A; Opsahl and others, 2018). 

These two samples were collected from Parkwood well in 

April 2016 during a period of rapid groundwater recharge, 

and the elevated atrazine concentrations measured in these 

samples were consistent with an influx of recently applied 

atrazine in recharge.

Factors Affecting Pesticides

The presence of trace concentrations of pesticides in 

rainfall in this study is consistent with concentrations of 

pesticides in rainfall in both developed and pristine areas of 

the United States (Thurman and Cromwell, 2000; Potter and 

Coffin, 2017). Postapplication losses of pesticides are reported 

to range from 0.2 to more than 90 percent and are related to 

many factors, including physical properties of the pesticide 

and climatic conditions (Van Pul and others, 1999; Barbash, 

2007). For example, volatilization can be a major dissipation 
Table 11. Summary of pesticide detections in rainfall, surface-water, and groundwater samples collected from sites in the San Antonio segment of tcentral Texas, 2012–16. [Estimated pesticide values reported at less than the laboratory reporting level were considered detections. NA, not applicable, Id, insecticide degradate; DEA, deethylatrazine; Cisopropylamino-6-amino-s-triazine; DCPA (dacthal), dimethyl tetrachloroterephthalate; Hd, herbicide degradate; H, herbicide; I, insecticide; Id, insecticide degradate]

Type of 

pesticide

Rainfall sites 

(fig 1; table 1)

Surface-water sites 

(fig. 1; table 1)

Groundwater site

(fig. 1; table 1) 

Rural 

rainfall1 Urban rainfall Frio River Seco Creek Culebra Creek Helotes Creek Seco well2 Parkwood well2 Zarzam

wel

Number of samples (n) NA 4 7 27 2 2 7 29 32 27 Pesticide compound Number of detections

1-Naphthol Id 0 0 0 0 0 0 0 0 1 

DEA (CIAT) Hd 1 4 4 1 0 3 16 32 26 Acetochlor H 1 0 0 0 0 0 0 0 0 Atrazine H 3 4 4 1 0 5 1 32 26 Azinphos-methyl I 0 0 0 0 0 0 0 0 1 Carbaryl I 0 3 2 0 0 3 0 0 0 

Chlorpyrifos I 2 0 1 0 0 0 0 0 0 Cypermethrin I 0 0 1 0 0 0 0 0 0 DCPA (dacthal) H 2 6 2 0 2 4 0 0 0 

Desulfinylfipronil amide Id 0 0 1 0 0 0 0 0 0 

Desulfinylfipronil Id 0 0 2 0 0 1 0 0 0 Dieldrin I 0 0 0 0 0 0 0 1 0 Fipronil sulfide Id 0 0 1 0 0 0 0 0 0 Fipronil sulfone Id 0 0 0 0 0 0 0 1 0 Fipronil I 0 0 2 0 0 0 0 0 0 Fonofos I 0 0 1 0 0 0 0 0 0 Malaoxon I 0 1 0 0 0 0 0 0 0 

Metolachlor H 4 3 8 2 1 1 0 1 0 

Prometon H 0 0 3 0 0 2 0 9 0 

Prometryn H 0 0 1 0 0 0 0 0 0 

Simazine H 0 0 0 0 0 1 2 32 25 

Tebuconazole H 1 0 0 0 0 0 0 0 0 

Trifluralin H 0 0 1 0 0 0 0 0 0 

Total number of detections per site NA 14 21 34 4 3 20 19 108 79 Average number of detections per sample5 NA 3.5 3.0 1.7 2.0 1.5 2.9 0.7 3.4 2.

1Rainfall samples collected at U.S. Geological Survey streamflow-gaging station 08195000 Frio River at Concan, Texas (Frio River site), were considered representative of ru2Well represents unconfined aquifer conditions. 3Samples from the Comal 1 spring site at Comal Springs, a spring at the downgradient end of the regional flow path (fig. 1), were for the purpose of this report considered gro

of deeper, confined groundwater.

4Calculated for each compound as the total number of detections divided by the total number of samples (n = 164).5Calculated for each site as the total number of detections of all pesticides divided by the number of samples.
52 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Date

Dec. 2011

Dec. 2012Dec. 2013Dec. 2014

Dec. 2015

Dec. 2016

0

0.01

0.02

0.03

0.04

0.05

Metolachlor, in micrograms per liter 

1

Detections consist of values greater than or equal to the laboratory reporting level (LRL) and estimated values. A constituent concentration is considered estimated by the laboratory

when results are greater than the long-term method detection level (LT–MDL) and less than the LRL; that is, a detection is considered likely, but quantification is considered questionable.

The remark code of “E” (estimated) is assigned by the laboratory for these values (Childress and others, 1999).

2

Nondetections are values reported for constituents that were analyzed for and were present at concentrations below the LRL and not quantified or were not present at all. For display

purposes only, nondetections are depicted on this graph at the concentration that defines the LRL.

Detection,1

 by site (fig. 1; table 1)

Rural rainfall—Sample collected at Frio

 River

Urban rainfall—Sample collected at 

 Urban rainfall 1, Urban rainfall 2, and

 Urban rainfall 3

Frio River

Seco Creek

Culebra Creek

Helotes Creek

Nondetection2

 (ND), by site (fig. 1; 

 table 1)

Urban rainfall ND—Sample collected at

 Urban rainfall 1, Urban rainfall 2, and

 Urban rainfall 3

Frio River ND

Culebra Creek ND

Helotes Creek ND

EXPLANATION

Figure 17. Concentrations of the herbicide metolachlor measured in surface-water and rainfall samples collected from sites in the San 

Antonio segment of the Edwards aquifer, south-central Texas, 2012–16. 

Metolachlor

DCPA

Carbaryl

DEA

Atrazine

Pesticide

0

0.01

0.02

0.03

0.04

0.05

0.06

Concentration, in micrograms per liter

DEA, deethylatrazine; DCPA (dacthal), dimethyl tetrachloroterephthalate 

EXPLANATION

Rainfall

Surface water

Figure 18. Comparison of the maximum concentrations of the five pesticides that were measured in both rainfall and surface-water 

samples collected from sites in the San Antonio segment of the Edwards aquifer, south-central Texas, 2012–16.
Pesticide Geochemistry 53

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

0

0.002

0.004

0.006

0.008

0.010

0.012

0.014

DEA, in micrograms per liter

760

780

800

820

840

Daily median water-level altitude, in feet

above North American Vertical Datum of 1988

EXPLANATION

Daily median water-level

altitude

DEA

DEA-ND

Seco well (fig. 1; table 1) Dry period1 Wet period2

DEA, deethylatrazine; ND, nondetection

Detections consist of values greater than or equal to the laboratory reporting level (LRL) and estimated values. A constituent concentration is considered estimated by the laboratory

when results are greater than the long-term method detection level (LT–MDL) and less than the LRL; that is, a detection is considered likely, but quantification is considered questionable.

The remark code of “E” (estimated) is assigned by the laboratory for these values (Childress and others, 1999).

Nondetections are values reported for constituents that were analyzed for and were present at concentrations below the LRL and not quantified or were not present at all. For display

purposes only, nondetections are depicted on this graph at the concentration that defines the LRL.

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the “wet period.”

Figure 19. Daily median water-level altitude and concentrations of the atrazine degradate deethylatrazine (DEA) measured in water 

samples from Seco well in the San Antonio segment of the Edwards aquifer, south-central Texas, 2013–16. 

route for DCPA, particularly under warm conditions with high 

soil moisture levels (U.S. Environmental Protection Agency, 

1998). The volatility of DCPA likely accounts for its frequent 

occurrence and low concentrations (0.0015–0.0032 μg/L) in 

rainfall during this study (Opsahl and others, 2018). Rainfall 

composition can be highly variable among and within storms, 

and it is likely that rainfall samples collected during this study 

do not reflect the full variability of rainfall compositions. 

The presence of a pesticide in rainfall is not necessarily 

an indication of pesticide origin—pesticides released into the 

atmosphere can be transported distances of a few meters to as 

many as thousands of kilometers (Van Pul and others, 1999). 

Three pesticides (atrazine DCPA, and metolachlor) and the 

pesticide degradate DEA were detected in both rural and urban 

rainfall at similar concentrations; the sources might be local 

or might be hundreds of kilometers distant. The insecticide 

carbaryl was detected in three of seven samples of urban rainfall 

but no samples of rural rainfall; conversely, the insecticide 

chlorpyrifos was detected in two of four samples of rural 

rainfall but no samples of urban rainfall. The disparate settings 

might indicate localized sources of these pesticides to rainfall, 

although these pesticides were frequently detected in rainfall 

samples collected in locations across the United States (Vogel 

and others, 2008). 

Rainfall should be considered a source of many of the 

pesticides frequently detected in surface water. Among the 

pesticides most commonly detected in surface water (atrazine, 

DEA, carbaryl, DCPA, metolachlor, and prometon), all but 

prometon also were detected in at least one sample of rainfall 

(table 11). The maximum concentration measured in rainfall 

was higher than that measured in surface water for four of the 

five pesticides detected in both types of samples (fig. 18). In 

contrast, prometon and nine pesticides detected infrequently 

in surface water were not detected in rainfall, indicating 

that either they might have a local source in the watershed 

or the full range of pesticide occurrences in rainfall is not 

represented in the dataset.

The assemblage of pesticides detected in samples of 

surface water collected from the three rural streams (Frio 

River, Seco Creek, and Culebra Creek) differed from the 

assemblage detected in samples collected from the more 

urban Helotes Creek (table 11). However, this land-cover 

comparison is limited because only seven urban surface-water 

samples were collected, and they were collected from only one 

stream (Helotes Creek). The herbicide metolachlor was the 

pesticide most frequently detected in samples of rural surface 

water (11 of 31 samples; 35 percent), but was detected in 

only 1 sample of urban surface water (14 percent of samples). 
54 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Concentration, in micrograms per liter

Daily median water-level altitude, in feet above

North American Vertical Datum of 1988 (NAVD 88)

Concentration, in micrograms per liter

Daily median water-level altitude, in

feet above NAVD 88

0.001

0.01

0.1

1

670

690

710

730

750

0

0.005

0.010

0.015

0.020

630

650

670

690

710

Dry period1 Wet period2

Dry period1 Wet period2

DEA, deethylatrazine; ND, nondetection

DEA, deethylatrazine; ND, nondetection

Detections consist of values greater than or equal to the laboratory reporting level (LRL) and estimated values. A constituent concentration is considered estimated by the laboratory

when results are greater than the long-term method detection level (LT–MDL) and less than the LRL; that is, a detection is considered likely, but quantification is considered questionable.

The remark code of “E” (estimated) is assigned by the laboratory for these values (Childress and others, 1999).

Nondetections are values reported for constituents that were analyzed for and were present at concentrations below the LRL and not quantified or were not present at all. For display

purposes only, nondetections are depicted on this graph at the concentration that defines the LRL.

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the “wet period.”

EXPLANATION

Daily median water-level

altitude

DEA

Atrazine

Simazine

Prometon

Prometon ND

EXPLANATION

Daily median water-level

altitude

DEA

Atrazine

Simazine

DEA ND

Atrazine ND

Simazine ND

Prometon ND

A. Parkwood well (fig. 1; table 1)

B. Zarzamora well (fig. 1; table 1)

Figure 20. Daily median water-level altitude and concentrations of deethylatrazine (DEA), atrazine, and simazine measured in water 

samples from urban wells including A, Parkwood well and B, Zarzamora well in the San Antonio segment of the Edwards aquifer, southcentral Texas, 2013–16.
Pesticide Geochemistry 55

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Concentration, in micrograms per liter

Comal Springs daily mean discharge,

in cubic feet per second

50

150

250

350

450

0

0.002

0.004

0.006

0.008

0.010

0.012

0.014

Dry period1 Wet period2

DEA, deethylatrazine; ND, nondetection

Detections consist of values greater than or equal to the laboratory reporting level (LRL) and estimated values. A constituent concentration is considered estimated by the laboratory

when results are greater than the long-term method detection level (LT–MDL) and less than the LRL; that is, a detection is considered likely, but quantification is considered questionable.

The remark code of “E” (estimated) is assigned by the laboratory for these values (Childress and others, 1999).

Nondetections are values reported for constituents that were analyzed for and were present at concentrations below the LRL and not quantified or were not present at all. For display

purposes only, nondetections are depicted on this graph at the concentration that defines the LRL.

EXPLANATION

Comal Springs daily mean

discharge

Sample collected from

Comal 1 spring

DEA

Atrazine

Simazine

Prometon

DEA ND

Atrazine ND

Simazine ND

Prometon ND

Comal 1 spring (fig. 1; table 1)

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the “wet period.”

Figure 21. Daily mean discharge at Comal Springs and concentrations of deethylatrazine (DEA), atrazine, simazine, and prometon 

measured in water samples collected from Comal 1 spring in the San Antonio segment of the Edwards aquifer, south-central Texas, 

2013–16. 

In contrast, the herbicide atrazine was the pesticide most 

frequently detected in samples of urban surface water (5 of 

7 samples; 71 percent), but was detected less frequently in 

samples of rural surface water (5 of 31 samples; 16 percent). 

Eight insecticides (5 parent pesticides and 3 degradates) were 

detected in at least 1 sample of rural surface water, whereas 

only 1 insecticide (carbaryl) and 1 insecticide degradate 

(desulfinylfipronil) were detected in urban surface water. 

The larger number of insecticides detected in rural than urban 

settings is the reverse of the general pattern in the United 

States reported by Gilliom and others (2006), although the 

data analyzed in that study were collected between 1992 and 

2001, and patterns might have changed over time.

Although surface water provides recharge to the aquifer, 

the assemblage of pesticides detected in groundwater was 

much less diverse than that in surface water (table 11). 

Pesticide detections in groundwater were dominated by 

atrazine, DEA, and simazine, which were among the 

five pesticides detected most frequently in a nationwide 

groundwater study during 1992–2001 (Gilliom and others, 

2006). The frequent occurrence of these pesticides in 

groundwater reflects their mobility and persistence. For 

example, the more frequent occurrence of metolachlor than 

atrazine in rural surface water but the relative absence of 

metolachlor in groundwater likely results from the more 

rapid breakdown of metolachlor than atrazine based on 

observations of breakdown rates in soils (soil half-lives of 

26 days for metolachlor and 146 days for atrazine) (Barbash 

and others, 1999). Dilution might also affect pesticide 

occurrence in Edwards aquifer groundwater; atrazine and 

DEA, when detected in surface water, occurred at higher 

concentrations (on the basis of their mean and maximum 

values) than did other pesticides (Opsahl and others, 2018). 

Their subsequent dilution within the aquifer therefore might 

still result in detectable concentrations.

Samples of urban surface water collected for this study 

are not fully representative of urban water in the San Antonio 

area. The comparison of urban surface water and groundwater 

is hampered by the small number of urban surface-water 

samples (seven) and the single urban surface-water site. 
56 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Date

July 2013

Jan. 2014

July 2014

Jan. 2015

July 2015

Jan. 2016

July 2016

Jan. 2017

Daily median water-level altitude, in feet above

North American Vertical Datum of 1988 (NAVD 88)

Daily median water-level altitude, in feet

above NAVD 88

0.001

0.01

0.1

1

0.001

0.01

0.1

1

670

690

710

730

750

670

690

710

730

750

Atrazine, in micrograms per liter Simazine, in micrograms per liter

Dry period1 Wet period2

Dry period1 Wet period2

Detections consist of values greater than or equal to the laboratory reporting level (LRL) and estimated values. A constituent concentration is considered estimated by the laboratory

when results are greater than the long-term method detection level (LT–MDL) and less than the LRL; that is, a detection is considered likely, but quantification is considered questionable.

The remark code of “E” (estimated) is assigned by the laboratory for these values (Childress and others, 1999).

Nondetections are values reported for constituents that were analyzed for and were present at concentrations below the LRL and not quantified or were not present at all. For display

purposes only, nondetections are depicted on this graph at the concentration that defines the LRL.

Daily median water-level

altitude from

Parkwood well

Sample, by site

(fig. 1; table 1)

Parkwood well

Zarzamora well

Zarzamora well ND

EXPLANATION

ND, nondetection

1

July 2013 through May 2015 is referred to as the “dry period.”

2

June 2015 through December 2016 is referred to as the “wet period.”

A

B

Figure 22. Daily median water-level altitude at Parkwood well and pesticide concentrations measured in groundwater samples 

collected from urban wells (Parkwood and Zarzamora wells) for A, atrazine and B, simazine in the San Antonio segment of the Edwards 

aquifer, south-central Texas, 2013–16. 
Pesticide Geochemistry 57

This lack of representativeness is evidenced by detections 

of the herbicide simazine, which was detected in only one 

sample of surface water but was consistently detected in 

shallow urban groundwater. However, simazine was detected 

in 23 percent of 40 samples collected from 5 urban recharge 

streams sampled between 2001 and 2012 (Opsahl, 2012). 

Additionally, Opsahl (2012) reported detections of the 

pesticides benfluralin, DCPA, diazinon, disulfoton, malathion, 

and pendimethalin, all of which were not detected in samples 

collected in this study. Other urban streams, not sampled for 

this study, likely contribute a greater diversity of pesticides 

to the aquifer. 

Land cover and aquifer hydrogeology influenced 

pesticide contamination of Edwards aquifer groundwater. 

Average number of pesticide detections for samples of 

groundwater from both the shallow, unconfined and deeper, 

confined urban wells (Parkwood and Zarzamora wells, 

respectively) were more contaminated by pesticides than were 

samples from the unconfined and rural Seco well (table 11). 

Atrazine, DEA, and simazine were detected consistently 

(more than 90 percent of samples) throughout the study in 

samples collected at the urban wells, and prometon was 

detected primarily during the dry period in the shallow, 

unconfined urban Parkwood well (fig. 20). At the rural Seco 

well, which is similar in depth to the deep, confined urban well 

(Zarzamora well), only DEA was detected, and only during 

the dry period when WLAs were low (fig. 19). In groundwater 

samples collected during 1996–2006, the most frequently 

detected pesticides (atrazine, DEA, and simazine) were 

detected in more than 50 percent of samples of shallow urban 

unconfined groundwater (Musgrove and others, 2010). Based 

on the high rate of detection of atrazine, DEA, and simazine in 

shallow urban unconfined groundwater, Musgrove and others 

(2010) concluded that “the urban environment is a source of 

anthropogenic contaminants to Edwards aquifer groundwater,” 

a finding that is supported by the results of this study.

Hydrogeological characteristics affected concentrations 

and temporal patterns of pesticides detected in the urban wells. 

Concentrations of atrazine were higher and more variable in 

the shallow, unconfined well (Parkwood) than at the deeper, 

confined well (Zarzamora); concentrations of simazine in the 

two wells were similar during the dry period, when WLAs 

were low, but diverged during the wet period, after WLAs 

began to rise (fig. 22A, B). In the urban wells, the ubiquity 

and relatively constant, low concentrations of atrazine, DEA, 

simazine (fig. 20A, B), and—in the shallow Parkwood well—

prometon during the dry period (fig. 20A) likely indicate 

legacy contamination present at trace concentrations in the 

diffuse flow of the aquifer. During the wet period, which was 

punctuated by at least three recharge events and rising WLAs, 

there was a dynamic response at the shallow Parkwood well 

(fig. 20A). Concentrations of atrazine and simazine peaked 

in May and early November 2015 and in mid-April 2016, 

coinciding with rapid decreases in SC (fig. 5B) that signaled 

an influx of recharging surface water. The highest simazine 

concentration occurred in November 2015, and the highest 

atrazine concentration occurred in April 2016; in each case, 

the concentration was about 20 times greater than the ambient 

concentration (figs. 20A, 22A, B). The higher concentrations 

measured during periods of groundwater recharge indicate 

that groundwater in the shallow Parkwood well is in close 

communication with surface water and therefore vulnerable 

to contamination. Although much more muted, there also 

was a slight response at the deep Zarzamora well: atrazine 

concentration increased slightly in May 2015, when the rise 

in the WLA first began as hydrologic conditions transitioned 

from dry to wet.

The geochemical time series at Comal 1 spring, which 

integrates flow from the entire aquifer segment, showed 

little response to the change in hydrologic conditions from 

the dry period to the wet period (fig. 21). Atrazine, DEA, 

simazine, and prometon were frequently detected, but at 

concentrations that were consistently low and similar to 

ambient concentrations in the urban wells; only prometon 

showed any systematic variability. The relatively unvarying 

concentrations indicate that the regional component of 

groundwater is buffered from brief spikes in contamination 

associated with recharge events, but also indicate that the 

presence of pesticides has become a ubiquitous feature of 

Edwards aquifer groundwater.

Pesticide Inputs and Cycling in the

Edwards Aquifer

The varied inputs of pesticides in rainfall, surface water, 

and groundwater in the Edwards aquifer region are evidence of 

anthropogenic sources of aquifer contamination and therefore 

important to current and future water-quality concerns for the 

Edwards aquifer. Given the small number of rainfall samples 

collected for this study and the large amount of variability in 

pesticide assemblage among rainfall events, rainfall samples 

likely do not represent the full range of variability across 

the entire study area. Similarly, for urban surface water, all 

urban samples are from one stream—Helotes Creek—results 

from which likely do not fully characterize the variability 

of potential recharge from streams in the urban San Antonio 

area. Nevertheless, the results of this study lead to several 

conclusions. The urban environment is a source of pesticide 

contamination to Edwards aquifer groundwater; urban wells, 

both shallow and deep, had higher detections of pesticides 

than did the rural well (table 11). The response to recharge 

events at the shallow, urban Parkwood well (fig. 5B) indicates 

that, in addition to background concentrations in diffuse flow, 

some pesticides are entering the aquifer with recharge water—

higher concentrations were observed in response to specific 

recharge events. Conversely, dilution by recharge water could 

decrease concentrations of other pesticides in groundwater; 

the trace concentrations of several pesticides not present 

in stormwater runoff during the dry period may have been 

diluted to less than the detection level during the wet period, 

as seen for Seco well (fig. 19). Those pesticides with higher 
58 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

concentrations are still likely to be detected, even with dilution 

that occurs in the aquifer.

Natural processes, such as biological degradation and 

alteration of individual pesticides, potentially affect the fate 

and transport of pesticides throughout the aquifer (Gilliom 

and others, 2006). Degradation of some pesticides, such as 

the degradation of atrazine to DEA, requires oxic conditions 

(Scribner and others, 2005), which occur in the Edwards 

aquifer. The presence of DEA in rainfall and surface water 

and their low DAR ratios (median DAR of 0.53 and 0.75 

for rainfall and surface water, respectively) in comparison 

to groundwater (median DAR of 1.86) (Opsahl and others, 

2018) indicate that continued atrazine degradation occurs 

during the transition from surface water to groundwater. 

The routine detection of the parent compound atrazine in 

groundwater samples, including discharge at Comal 1 spring 

(fig. 21), however, indicates that groundwater residence times, 

traveltimes, and other factors do not allow for the complete 

degradation of atrazine within the aquifer.

Factors Affecting Aquifer Vulnerability 

The Edwards aquifer, like many karst aquifers, is highly 

susceptible to contamination because of high porosity and 

permeability and rapid flow of recharging surface water 

through conduits. Measured concentrations of NO3 and 

pesticides—the compounds of focus for this study—are 

currently low relative to levels of regulation for drinking 

water (for example, 10 mg/L for NO3

 and 3 μg/L for atrazine; 

U.S. Environmental Protection Agency, 2017b). Nonetheless, 

elevated NO3

 concentrations relative to national background 

values (Dubrovsky and others, 2010) and the widespread 

detection of pesticides indicate that the Edwards aquifer is 

vulnerable to contamination and that vulnerability is affected 

by factors such as land cover, aquifer hydrogeology, and 

changes in hydrologic conditions.

Land Cover

Land cover has a strong influence on the vulnerability of 

Edwards aquifer groundwater to the occurrence of nutrients 

and pesticides. Greater vulnerability of groundwater in 

urban areas relative to rural areas was evident from results 

for urban groundwater sites, which generally had higher 

NO3 concentrations, elevated δ15N-NO3

 values, a greater 

diversity of pesticide detections, and higher pesticide 

concentrations (fig. 22; tables 9, 11). The continuum of water 

quality from unconfined rural groundwater (least affected 

by anthropogenic contamination) to unconfined urban 

groundwater (most affected by anthropogenic contamination) 

sites (fig. 14) demonstrates enhanced vulnerability of 

urban versus rural land cover. The deeper, confined urban 

groundwater sites were generally more similar to the urban 

unconfined sites than to the rural unconfined site with 

respect to the occurrence and concentration of nutrients and 

pesticides. There were, however, spatial variations in the 

anthropogenic contributions to the shallow, urban unconfined 

wells; for example, Parkwood well had a greater degree of 

NO3

 contamination than did Shavano well. Differences in 

contaminant occurrences and concentration among unconfined 

urban wells indicate that the urban parts of the aquifer are 

not uniformly vulnerable, but rather are affected by spatial 

differences in the sources of nutrients and pesticides.

Aquifer Hydrogeology

Comparison of groundwater samples collected from 

the shallow, unconfined and deeper, confined parts of the 

aquifer indicates that aquifer hydrogeology—specifically 

including aquifer confinement and recharge pathways—

affects groundwater vulnerability. In urban areas, the shallow, 

unconfined groundwater sites showed greater temporal 

variability in both nutrient and pesticide concentrations, as 

well as a greater degree of contamination, than did deeper, 

confined groundwater sites. In comparison to that of the 

shallow, unconfined groundwater sites, the water quality of 

the deeper, confined groundwater sites was relatively invariant 

during this multiyear study. Tracers of water-rock interaction 

indicate that confined groundwater is well mixed and more 

geochemically evolved with longer residence times (fig. 7); 

however, most confined groundwater sites had elevated NO3

concentrations and exhibited pesticide detections, providing 

evidence for a component of local recharge. The presence 

of a component of local recharge is consistent with previous 

studies that have indicated that, while groundwater throughout 

the Edwards aquifer is generally young (with mixtures ranging 

from less than 1 to about 40 years in apparent age), deeper 

groundwater is older and less influenced by anthropogenic 

activities (Musgrove and others, 2011). In this study, the 

relation between the degree of confinement and influence 

from anthropogenic activities is demonstrated by samples 

from the urban, confined Mission well, which had the highest 

Mg/Ca ratios, evidence of groundwater with longer residence 

time that has not—to date—been substantially affected by 

anthropogenic activities. Continued loading of compounds 

of concern in urban recharge, however, might ultimately 

affect deeper parts of the aquifer as legacy contamination 

accumulates in diffuse flow. Although aquifer hydrogeology 

is an important factor related to aquifer vulnerability, land 

cover likely has a greater influence on pesticide contamination 

of groundwater.

Changes in Hydrologic Conditions

Changes in hydrologic conditions affect water quality 

and the associated transport of nutrients and pesticides by 

providing conditions conducive for transport or dilution in the 

aquifer, particularly during periods of groundwater recharge. 

Temporal variability in hydrologic conditions for the Edwards 
Summary 59

aquifer is apparent in data for surface water as a source of 

groundwater recharge (figs. 6, 7), WLA in wells (fig. 5A–F), 

spring discharge (fig. 5G), and groundwater quality (figs. 7–9, 

12–22). The observed temporal variability affects recharge 

sources, recharge amounts, groundwater traveltimes, flow 

routing, water-rock interaction processes, dilution, mixing, 

and, in turn, water quality. Observed variation in nutrient and 

pesticide concentrations at streams in response to changing 

flow conditions (figs. 10–11; Opsahl and others, 2018) 

indicates that recharge water quality is variable over time and 

may be different for each event. Additionally, the variability 

of water quality observed at shallow, unconfined wells (for 

example, Parkwood and Shavano wells) from the dry period 

to the wet period, when water quality becomes more variable, 

and at Parkwood well in response to recharge events, indicates 

that water quality is dynamic on multiple time scales. An 

understanding of the relation between variability in hydrologic 

conditions and water quality provides a foundation on which 

to assess long-term changes in water quality and to distinguish 

between short-term, hydrologically induced variation and 

longer term trends. 

Summary

The San Antonio segment of the Edwards aquifer, located 

in south-central Texas, is a designated sole-source aquifer 

and the primary water supply for more than 1.7 million 

people in the City of San Antonio and surrounding area. The 

Edwards aquifer recharge zone is particularly susceptible to 

contamination from the land surface because of the dynamic 

and rapid recharge that occurs through karst features such as 

sinkholes and losing streams during groundwater recharge 

events. The U.S. Geological Survey, in cooperation with the 

San Antonio Water System, performed a study from 2010 to 

2016 to better understand how water quality changes under 

a range of hydrologic conditions and in contrasting landcover settings (rural and urban) in the Edwards aquifer. The 

study design included continuous hydrologic monitoring, 

continuous water-quality monitoring, and discrete sample 

collection for a detailed characterization of water quality at 

a network of sites throughout the aquifer system. The sites 

were selected to encompass a “source-to-sink” (that is, from 

aquifer recharge to aquifer discharge) approach. Network 

sites were selected to characterize rainfall, recharging surface 

water, and groundwater; groundwater sites included wells 

in the unconfined part of the aquifer (unconfined wells) and 

in the confined part of the aquifer (confined wells) and a 

major discharging spring. Storm-related samples—including 

rainfall samples, stormwater-runoff (surface-water) samples, 

and groundwater samples—were collected to characterize the 

aquifer response to recharge.

Rainfall often contained measurable concentrations 

of nitrate, nitrite, dissolved organic nitrogen, and 

orthophosphate at both rural and urban rainfall sites, indicating 

that atmospheric deposition is a source of nitrogen and 

phosphorous to the study area. Nitrate was the dominant 

nitrogen species at most surface-water sites. The median 

nitrate concentrations at surface-water sites were similar to or 

higher than the estimated national background concentration 

for streams. The higher concentration at Helotes Creek might 

result from a contribution of nitrate from urban sources. 

Median nitrate concentrations for both rural and urban 

surface-water sites examined in this study were substantially 

lower than the national background concentration for streams 

in agricultural areas and are therefore consistent with minor 

inputs of fertilizer.

The total variability in nitrate at Seco well was relatively 

small. Although the difference in nitrate concentration 

between the dry period and the wet period was small, it is 

distinguishable outside of the nitrate sensor’s precision, 

indicating that nitrate concentrations at this site responded 

to longer term changes in hydrologic conditions. In response 

to frequent rainfall during the transition to the wet period 

in spring 2015, water-level altitude at Parkwood well rose 

rapidly, by about 40 feet over a period of a few days, and 

concurrently, nitrate concentration decreased from about 2.7 

milligrams per liter to less than 1 milligram per liter. Mixing 

between ambient groundwater and lower nitrate recharging 

stormwater runoff derived directly from recharge streams such 

as Helotes Creek can account for the rapid and short-term 

dilution of ambient water at Parkwood well.

The small range of delta nitrogen-15 of nitrate (δ15Nnitrate [NO3

]) values for Seco well and corresponding low 

NO3

 concentrations indicate predominantly natural, soilderived NO3

 with little contribution from anthropogenic 

sources. In contrast, Parkwood well (with the exception of 

samples collected in response to recharge events) had the 

highest δ15N-NO3

 values and NO3 concentrations, which 

is consistent with mixing with a component of NO3

 from 

human or animal waste sources. Other groundwater sites 

were intermediate in their δ15N-NO3

 and concentration values, 

with higher δ15N-NO3

 values corresponding to higher NO3

concentrations. NO3

 concentrations at Parkwood well indicate 

that about two-thirds of the NO3

 in groundwater samples 

collected for this study might result from anthropogenic NO3

sources, assuming that 1 milligram per liter is the natural 

background concentration and the remainder (2.12 milligrams 

per liter) is anthropogenic.

Natural biogeochemical processes in the aquifer affect 

the fate and transport of all nitrogen species and play a role in 

determining the distribution of NO3

 concentrations throughout 

the aquifer. The process of nitrification of dissolved organic 

nitrogen and ammonia from rainwater and surface water 

might be an important but previously unaccounted for source 

of NO3

 to Edwards aquifer groundwater, as evidenced by 

the decrease in dissolved organic nitrogen plus ammonia 

concentrations from rainfall to surface water to groundwater 

and the corresponding increase in NO3

 concentration. 

Characterization of sources, fate, and transport of NO3

 in 

the San Antonio segment of the Edwards aquifer therefore 

requires consideration of all nitrogen species.
60 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

At least 1 pesticide was detected in each of the 11 rainfall 

samples analyzed. Sixteen pesticides were detected in at least 

1 surface-water sample. Nine pesticides were detected in at 

least 1 of the 115 groundwater samples collected. Among the 

unconfined wells, pesticides occurred much less frequently 

at the rural Seco well than at the 2 urban wells (Parkwood 

and Zarzamora). Average number of pesticide detections for 

samples of groundwater from both the shallow, unconfined 

and deeper, confined urban wells (Parkwood and Zarzamora, 

respectively) were more contaminated by pesticides than were 

samples from the unconfined and rural Seco well.

The ubiquity and relatively constant, low concentrations 

of atrazine, deethylatrazine, simazine, and—in the shallow 

Parkwood well—prometon during the dry period likely 

indicate legacy contamination present at trace concentrations 

in the diffuse flow of the aquifer. The relatively unvarying 

concentrations indicate that the regional component of 

groundwater is buffered from brief spikes in contamination 

associated with recharge events, but also indicate that the 

presence of pesticides has become a ubiquitous feature of 

Edwards aquifer groundwater. The response to recharge 

events at the shallow urban Parkwood well indicates that, 

in addition to background concentrations in diffuse flow, 

some pesticides are entering the aquifer with recharge 

water—higher concentrations were observed in response 

to specific recharge events. The routine detection of 

the parent compound atrazine in groundwater samples, 

including discharge at Comal 1 spring, however, indicates 

that groundwater residence times, traveltimes, and other 

factors do not allow for the complete degradation of atrazine 

within the aquifer.

Elevated NO3 concentrations relative to national 

background values and the widespread detection of 

pesticides indicate that the Edwards aquifer is vulnerable to 

contamination and that vulnerability is affected by factors 

such as land cover, aquifer hydrogeology, and changes in 

hydrologic conditions. Greater vulnerability of groundwater 

in urban areas relative to rural areas was evident from 

results for urban groundwater sites, which generally had 

higher NO3

 concentrations, elevated δ15N-NO3 values, 

a greater diversity of pesticides, and higher pesticide 

concentrations. Differences in contaminant occurrences and 

concentration among unconfined urban wells indicate that 

the urban parts of the aquifer are not uniformly vulnerable, 

but rather are affected by spatial differences in the sources 

of nutrients and pesticides. In urban areas, the shallow, 

unconfined groundwater sites showed greater temporal 

variability in both nutrient and pesticide concentrations, as 

well as a greater degree of contamination, than did deeper, 

confined groundwater sites. In comparison to that of the 

shallow, unconfined groundwater sites, the water quality 

of the deeper, confined groundwater sites was relatively 

invariant during this multiyear study. Continued loading 

of compounds of concern in urban recharge, however, 

might ultimately affect deeper parts of the aquifer as legacy 

contamination accumulates in diffuse flow. Although aquifer 

hydrogeology is an important factor related to aquifer 

vulnerability, land cover likely has a greater influence 

on pesticide contamination of groundwater. Temporal 

variability in hydrologic conditions for the Edwards 

aquifer is apparent in data for surface water as a source of 

groundwater recharge, water-level altitude in wells, spring 

discharge, and groundwater quality. Observed variation in 

nutrient and pesticide concentrations at streams (for example, 

the Frio River and Helotes Creek) in response to changing 

flow conditions indicates that recharge water quality is 

variable over time and may be different for each event. 

Additionally, the variability of water quality observed at 

shallow, unconfined wells (for example, Parkwood and 

Shavano) from the dry period to the wet period, when water 

quality becomes more variable, and at Parkwood well in 

response to recharge events indicates that water quality is 

dynamic on multiple timescales. An understanding of the 

relation between variability in hydrologic conditions and 

water quality provides a foundation on which to assess longterm changes in water quality and to distinguish between 

short-term, hydrologically induced variation and longer 

term trends.

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68 Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas

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Opsahl and others—Water-Quality Observations of the San Antonio Segment of the Edwards Aquifer, Texas—SIR 2018–5060

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