Purpose and Scope

The purpose of this report is to document the results of a synoptic sampling done by the USGS in cooperation with the NPGCD during 2019–20 and compare the results to those of a previous synoptic sampling conducted during 2012–13 (Baldys and others, 2014). Depth to water-level measurements and groundwater-quality samples were collected from 30 monitoring wells completed in the Ogallala aquifer in the NPGCD management area as part of this synoptic sampling (fig. 1; table 1). Groundwater-quality samples were collected for the analysis of dissolved solids, major ions, nutrients, and trace elements; selected organic compounds also were analyzed in 6 of the 30 wells. Results from the 2019–20 sampling period for selected constituents were compared to U.S. Environmental Protection Agency (EPA) drinking-water standards and to the 2012–13 sampling results. The groundwater-quality data from the samples collected were also assessed in the context of results obtained from quality-control samples collected in conjunction with the environmental samples.

Description of Study Area

The study area for this report coincides with the area managed by the NPGCD, which consists of the five northernmost counties that form the border between Texas and Oklahoma (Dallam, Sherman, Hansford, Ochiltree, and Lipscomb Counties) and parts of Hartley, Moore, and Hutchinson Counties north of the Canadian River where the rocks that contain the Ogallala aquifer are present at land surface. (fig. 1).

During 1981–2010, the average annual precipitation for the 7,390-square-mile study area ranged from 15.7 inches (in.) at Dalhart, Texas (Dallam County), in the western part of the study area to 23.2 in. in Follett, Tex. (Lipscomb County), in the eastern part of the study area (National Oceanic and Atmospheric Administration, 2021). In the study area, the NPGCD estimates that average annual recharge rates are less than one-third of an inch per year (NPGCD, 2020). Most of this negligible amount of recharge is from precipitation infiltrating through streambeds and playas and eventually reaching the groundwater table (Scanlon and others, 2002; McGuire, 2017).

Geologic and Hydrogeologic Setting

In the study area, the Ogallala aquifer is contained in the Ogallala Formation, a sand and gravel deposit capped with caliche. The Ogallala Formation was deposited during the Miocene Epoch of the Tertiary Period (Gutentag and others, 1984; Ryder, 1996). The Ogallala Formation is the principal water-bearing geologic unit of the central High Plains aquifer (Gutentag and others, 1984; Ryder, 1996); historically, the High Plains aquifer system was often referred to as the “Ogallala aquifer” throughout its extent in reference to the predominant water-bearing unit of the aquifer system, the Ogallala Formation (Dugan and others, 1994). In places, the Ogallala aquifer is in hydraulic connection with permeable parts of the underlying bedrock, which ranges in age from Permian to Cretaceous (Ryder, 1996). Regional uplift and erosion have stripped away the Great Plains (Ogallala Formation) in many places, but a large central area was little affected by eroding streams and is preserved. This preserved remnant of the uplifted area that contains the Ogallala Formation is known as the High Plains. Although the land surface of the High Plains has been modified little by streams, it has been pitted by carbonate dissolution and deflation, thus forming many playas (shallow depressions) that collect and store water during periods of precipitation and runoff. (Ryder, 1996). The Canadian River has cut through much of the Ogallala Formation in the Texas Panhandle. The Ogallala Formation consists primarily of heterogeneous sequences of coarse-grained sand and gravel, which originated from alluvial fan deposits and fluvial, lacustrine, and eolian deposits (NPGCD, 2020; Ryder, 1996). The NPGCD reports that the average thickness of the Ogallala Formation in the study area ranges from 115 ft in Moore County to 225 ft in Lipscomb County, with an estimated study-area-wide average aquifer thickness of 148 ft (NPGCD, 2020).

Depth to Water Level Methods

The same methods used in the previous synoptic study (Baldys and others, 2014) by USGS personnel to make depth to water-level measurements and collect water-quality samples were used in this study. At each of the monitoring wells shown in figure 1, the depth to water was measured before collecting water-quality samples to help ensure the sampling hose intake was at the proper depth to provide a representative sample and to determine any changes in groundwater levels (fig. 1; Ging and Mobley, 2021). The depth to water was measured at most wells by using USGS-owned electric groundwater-level measurement tapes (electric tapes or e-tapes) in accordance with methods described in Cunningham and Schalk (2011). The electric tapes used by USGS personnel were calibrated before use with a steel reference tape. An electric tape commonly is accurate to plus or minus 0.01 ft and is most accurate for groundwater levels less than 200 ft below land surface (Cunningham and Schalk, 2011). Depths to water in two of the wells where the groundwater depth exceeded 500 ft were measured by USGS personnel with a NPGCD-owned electric tape that could measure groundwater depths as great as 750 ft below land surface; the electric tape belonging to NPGCD was not subjected to the same reference calibration checks as the USGS-owned electric tape that was used.

NPGCD county well identifiers assigned to each well were used to identify the wells in this report and typically consist of the first two letters of the county name followed by a sequence of four numbers ( Baldys and others, 2014; NPGCD, 2020; Ging and Mobley, 2021). The county well identifiers assigned by the NPGCD supplement the 15-digit USGS station numbers used to identify each well in the National Water Information System database (table 1; USGS, 2021).

Water-Quality Sample Collection and Processing

Water-quality samples were collected from each of the 30 wells and processed onsite. Quality-control samples consisting of blank samples, replicate samples, and spike samples were collected to assess bias and precision in sampling and processing procedures. Groundwater-quality samples were processed in a mobile water-quality laboratory by following procedures outlined in the USGS “National Field Manual for the Collection of Water-Quality Data” (USGS, variously dated). A submersible pump was lowered into each well and was attached to stainless steel pipe sections to pump the groundwater directly from the wells to a manifold, which was used to direct the water to polytetrafluoroethylene (PTFE) tubing line that was routed into the mobile water-quality laboratory for sample processing.

Each well was purged to ensure that samples collected for analyses were representative of the groundwater in the aquifer. During the purging process, selected physicochemical properties (dissolved oxygen, pH, specific conductance, groundwater temperature, and turbidity) were measured and recorded as the unfiltered groundwater was pumped from the well. Standard USGS procedures require monitoring these physicochemical properties by using a multisensor meter installed in an airtight flow chamber until readings for the physical properties stabilize (USGS, variously dated). Each well was purged of at least three well casing volumes of groundwater prior to sample collection.

Once the physicochemical properties had stabilized, the well discharge was redirected to an enclosed processing chamber in the mobile water-quality laboratory where groundwater samples were processed for analysis. Sample collection and preservation were completed in the processing chamber to prevent the introduction of airborne contaminants. The tubing and fittings used for sample collection and processing were made of stainless steel or PTFE. The bottles used for sample collection were made of baked glass or polypropylene. The bottles were rinsed with deionized water prior to sample collection. Sampling equipment was thoroughly cleaned after each well was sampled following standard USGS procedures (USGS, variously dated). The stainless-steel pipe sections used to lower the submersible pump into the well were washed between sites with a pressurized water spray and then rinsed with deionized water (a high-volume rinse). Pipe sections were capped at each end during transport to prevent the introduction of dust and other airborne contaminants. Samples analyzed for major ions, trace elements, and nutrients were filtered through a 0.45-micrometer (μm) capsule filter. Pesticide samples analyzed for pesticide and herbicide compounds were collected from 5 of the 30 wells during 2019 and from 1 well in 2020; pesticide and herbicide compounds frequently detected in previous nationwide studies were targeted for sampling (Barbash and Resek, 1996; Kolpin and others, 1998; Scribner and others, 2007). All of the pesticide samples were filtered through a 47-millimeter (mm) diameter glass fiber filter with a pore size of 0.70 μm. Samples that required refrigeration (nutrients and organic compounds) were kept chilled at 4 degrees Celsius (°C) and shipped chilled to the appropriate laboratories for analyses.

Laboratory Analysis

Groundwater-quality samples were analyzed for major ions, nutrients, and trace elements at the National Water Quality Laboratory (NWQL), in Lakewood, Colorado, and for selected organic compounds at the NWQL or at the Organic Geochemistry Research Laboratory (OGRL), in Lawrence, Kansas. Organic compound samples were submitted to the NWQL for analyses of 81 common pesticides and to the OGRL for analyses of 35 common pesticides. Descriptions of laboratory analytical methods are published in Fishman and Friedman (1989) and Fishman (1993) for major ions analyses; Patton and Kryskalla (2003) for nutrients analyses; and Fishman and Friedman (1989), Struzeski and others (1996), Garbarino (1999), and Garbarino and others (2006) for trace-element analyses. Pesticide analysis was done by solid-phase extraction and gas chromatography/mass spectrometry following the methods published in Zaugg and others (1995), Lindley and others (1996), Sandstrom and others (2001), and Madsen and others (2002).

Concentrations of major ions, nutrients, and trace elements were reported by the NWQL and reviewed for completeness and accuracy. The analytical quantification procedure used by the NWQL for reporting results is based on the long-term method detection level (LT-MDL) and laboratory reporting level (LRL). The LT-MDL acts as a censoring limit for most analytical methods at the NWQL, and its purpose is to limit the false positive rate to less than or equal to 1 percent (Foreman and others, 2021). The LRL is twice the LT-MDL and was established to limit the false negative rate to less than or equal to 1 percent (Foreman and others, 2021). A constituent concentration is considered estimated by the laboratory when results are greater than the LT-MDL and less than the LRL; that is, detection is considered likely, but quantitation is considered estimated. Therefore, a remark code of “E” (estimated) is assigned by the laboratory for these results. Additionally, when measured values were greater than the calibration range of the instrument or when the field measurement stabilization criteria were not met, the data are reported with an “E” remark code (Foreman and others, 2021). In rare instances, a value lower than the LT-MDL might be reported at the discretion of the analyst. In contrast to the LT-MDL and LRL reporting convention used by the NWQL, the OGRL uses one reporting limit for each constituent that is set equal to or greater than the limit of quantitation (Meyer and others, 2009).

Quality-Assurance and Quality-Control Procedures

Quality-control data were collected to assess the variability and bias that may exist within the sample-collection procedures and laboratory analyses (USGS, variously dated). To test for this variability and bias, two field-blank samples, three sequential-replicate samples, and one matrix-spiked sample for organic compounds were collected in conjunction with environmental samples. Both field-blank samples were analyzed for inorganic constituents and organic compounds.

Hydrochemical Facies and Dissolved-Solids Concentrations

The composition of groundwater and its hydrochemical facies principally are controlled by the composition of recharge water, rock-water interaction, and the mixing of groundwater from different sources (Jones and others, 1997; Elango and Kannan, 2007). The term “hydrochemical facies” refers to a classification scheme used to describe water in terms of the major cation and anion milliequivalents composition. The trilinear diagram (Piper, 1944) is a useful tool for evaluating the relative abundance of major cations and anions and classifying hydrochemical facies or groundwater types, and it was used to evaluate groundwater samples collected from wells completed in the Ogallala aquifer in the Texas Panhandle during 2019–20 (fig. 3). Most of the groundwater samples collected during 2019–20 were classified as magnesium-bicarbonate type groundwater, which is similar to the classification of groundwater in samples collected during the 2012–13 (Baldys and others, 2014). Exceptions to the predominant magnesium-bicarbonate type classifications included two samples classified as sodium-bicarbonate type obtained from wells Da-3588 and Da-3589, the sample classified as sodium-chloride type obtained from well Oc-4465, and the sample classified as a mixed groundwater type obtained from well Hu-4855. Water samples obtained from wells Da-3588, Ds-3589, and Oc-4465 during 2019–20 were assigned the same groundwater types previously assigned when the wells were sampled during 2012–13 (Baldys and others, 2014). Water from well Hu-4855 seemed to have shifted from a calcium-chloride type groundwater during the 2012–13 sampling event to a more mixed water type, which included larger contributions of magnesium and bicarbonate, during the 2019–20 sampling event.

Dissolved-solids concentration is a measure of the quantity of dissolved constituents in a sample. Sodium and chloride are often the primary constituents that contribute to elevated dissolved-solids concentrations (Hem, 1985). As groundwater levels decline in an aquifer, deeper saline groundwater can move into areas being affected by groundwater withdrawals (Alley and others, 1999). Therefore, increasing sodium and chloride concentrations levels in groundwater can be a cause for concern among well owners. Dissolved-solids concentrations in samples for the wells sampled during the 2019–20 sampling period ranged from 260 to 774 mg/L with a median dissolved-solids concentration of 316 mg/L, which is slightly higher than the median dissolved-solids concentration for the 2012–13 sampling period of 311 mg/L (fig. 4; Ging and Mobley, 2021). Dissolved-solids concentrations in samples at five wells (Hu-4855, Oc-4465, Li-3687, Oc-4196, and Da-3589) were higher than the secondary drinking water standard of 500 mg/L (EPA, 2009) during the 2019–20 sampling period (fig. 4). Dissolved-solids concentrations in samples at four of those five wells were higher than the secondary drinking water standard of 500 mg/L in the 2012–13 sampling period as well (Ging and Mobley, 2021). Sodium concentrations in samples from the wells collected during the 2019–20 sampling period ranged from 7.5 to 200 mg/L with a median sodium concentration of 23.6 mg/L, which is slightly higher than the median sodium concentration of 21.2 mg/L for the 2012–13 sampling period (fig. 5; Ging and Mobley, 2021). Chloride concentrations measured in samples collected during the 2019–20 sampling period ranged from 4.0 to 253 mg/L with the median chloride concentration of 12.8 mg/L, which is slightly higher than the median chloride concentration of 11.6 mg/L for the 2012–13 sampling period (fig. 5; Ging and Mobley, 2021). Accounting for variability with analyses, median major ion concentrations were similar during the 2012–13 and 2019–20 sampling periods.

Nutrients

Nitrogen and phosphorus are usually the most commonly detected nutrients when elevated nutrient concentrations are found in groundwater (Ator and Denis, 1997). Ator and Denis (1997, p. 2) explain “although natural sources of nitrogen and phosphorus exist, elevated concentrations of these nutrients in groundwater are typically anthropogenic (human-derived).” Elevated nitrogen and phosphorus concentrations in groundwater frequently result from surface-water infiltration in agricultural or municipal areas where nutrient sources typically include livestock manure, septic systems, wastewater treatment systems, fertilizers, and atmospheric deposition of nitrogen from fossil fuel combustion (Ator and Denis, 1997; Dubrovsky and others, 2010). National background concentrations of nitrate, the most common form of nitrogen in groundwater, are estimated to be 1 mg/L (Dubrovsky and others, 2010). Water with nitrate concentrations higher than 10 mg/L can cause methemoglobinemia (also known as blue-baby syndrome) in infants less than 6 months of age and is considered a major contaminant and threat to groundwater quality in Texas (Chaudhuri and others, 2012). Samples were collected for the determination of four species of nitrogen and one species of phosphorus (orthophosphate). The four species of nitrogen that were analyzed were ammonia, nitrite plus nitrate, nitrite, and total nitrogen. Although organic nitrogen concentrations were not computed for this report, the concentration of organic nitrogen can be estimated by subtracting the sum of ammonia and nitrite plus nitrate from total nitrogen (Patton and Truitt, 2000; Ging and Mobley, 2021). Ammonia concentrations were less than the LRL of 0.010 mg/L for 27 of the 30 wells that were sampled during 2019–20. Nitrite concentrations were less than the LRL of 0.001 mg/L for all wells sampled during 2019–20 (Ging and Mobley, 2021). Nitrate is the dominant nitrogen species in the nitrate plus nitrite analysis because all nitrite concentrations were below the LRL. The median nitrate concentration for the 2019–20 sampling period was 2.25 mg/L, which is similar to the national median nitrate concentration of 2 mg/L determined by Mueller and Helsel (1996). However, nitrate concentrations in samples from two wells (Da-3688 and Sh-4136) exceeded the MCL of 10 mg/L (EPA, 2009) for drinking-water supplies during the 2019–20 sampling period (fig. 6). Wells Da-3688 and Sh-4136 were the same two wells where the nitrate concentrations exceeded 10 mg/L during the 2012–13 sampling period (Baldys and others, 2014). Nitrate concentrations were less than the LRL of 0.08 mg/L for two wells (Da-3588 and Da-3589), which are the only two wells where dissolved oxygen concentrations were less than 1 mg/L (Ging and Mobley, 2021). Total nitrogen concentrations were similar to the nitrate concentrations, indicating that nitrate concentrations account for most of the nitrogen concentrations (Ging and Mobley, 2021). Total nitrogen is defined as the sum of ammonia, nitrite, nitrate, and organic nitrogen (Patton and Truitt, 2000). When comparing median concentrations of nitrate and total nitrogen from the 2012–13 sampling period to the 2019–20 sampling period, median concentrations for both nitrate and total nitrogen increased slightly from the 2012–13 sampling period (2.05 and 2.14 mg/L, respectively) to the 2019–20 sampling period (2.25 and 2.31 mg/L, respectively) (fig. 7). Nitrate concentrations in samples collected from 19 of the 30 wells increased between the two sampling periods, the magnitudes of the increases ranged from 0.01 to 7.9 mg/L. Total nitrogen concentrations in samples collected from 17 of the 30 wells increased between the two sampling periods, and the magnitudes of the increases ranged from 0.01 to 9.4 mg/L (fig. 7; Ging and Mobley, 2021). Accounting for variability with analyses, median concentrations for nutrients species (both nitrogen and phosphorus) were similar for the 2012–13 and 2019–20 sampling periods.

Phosphorus is important in agricultural systems for the development of high-yielding crops, especially corn, soybeans, and wheat (Domagalski and Johnson, 2012). Long-term over-application of manure and chemical fertilizers that contain phosphorus can lead to phosphorus migration into the groundwater system, resulting in potential contamination of the groundwater resource. In water, phosphorous may be present as orthophosphate (Domagalski and Johnson, 2012). Because the median orthophosphate concentration was less than 0.019 mg/L in samples collected from the 30 wells sampled during 2019–20, it is unlikely that phosphorus contamination of groundwater from agricultural practices is common in the study area. The median orthophosphate concentration in samples collected increased slightly from 0.014 mg/L for the 2012–13 sampling period to 0.019 mg/L for the 2019–20 sampling period (fig. 7).

Trace Elements

Environmental samples from each well were analyzed to determine the concentrations of 23 trace elements, which are substances that typically are present in natural waters at concentrations less than 1.0 mg/L (Hem, 1985). Trace elements are present in small amounts in the environment and can originate from natural and manmade sources (Ayotte and others, 2011). Concentrations of trace elements are more likely to be higher in groundwater than in surface water as a result of rock-water interaction during the movement of groundwater through the rocks that compose the aquifers. Much of the groundwater in the Ogallala aquifer was recharged thousands of years ago (McMahon and others, 2007). Groundwater that has been in an aquifer for a long time can foster extensive rock-water interaction, which can cause trace element concentrations to increase (Elango and Kannan, 2007). Human activities such as fertilizer use, mining activities, and waste disposal also can affect concentrations of trace elements in groundwater (Ayotte and others, 2011). In water-quality samples collected from the 30 wells, concentrations of 16 trace elements were higher than the LRL in at least 50 percent of the samples. MCLs or secondary drinking-water standards are available for only 14 of the 23 trace elements analyzed (table 2). None of the trace element concentrations exceeded their respective MCLs or secondary drinking-water standards (Ging and Mobley, 2021). Median concentrations of trace elements for the 2012–13 sampling period were compared to those for the 2019–20 sampling period for constituents having at least 50 percent of detections in samples above the highest applicable LRL, and variability in analyses was accounted for. Comparison results indicated that median concentrations of two trace elements (lithium and uranium) increased and the median concentrations for two trace elements (barium and molybdenum) decreased (fig. 8). Arsenic concentrations, which can be a concern in the Ogallala aquifer because of agricultural use in the area, were all below the MCL of 10 μg/L in the 30 samples collected from all 30 wells sampled in 2019–20 (fig. 9; Ging and Mobley, 2021). Arsenic can naturally occur in rocks and soils but also was used in certain pesticide applications in the study area (Welch and others 2000; Thelin and Stone, 2010).

Organic Compounds

Detections of organic compounds such as pesticides in groundwater samples are potential indicators of anthropogenic effects on groundwater quality. Organic compounds detected in groundwater samples likely originate from surface water, which contains pesticides, seeping into the groundwater in the study area. Groundwater samples were submitted to the NWQL for analyses of 81 common pesticides and to the OGRL for analyses of 35 common pesticides. Samples were collected for organic compound analyses at 6 of the 30 wells sampled in the 2019–20 study (Ging and Mobley, 2021). Organic compounds were detected in samples collected from four of the six wells (Sh-2369, Hn-1917, Ha-2480, and Li-0656) (fig. 10). Atrazine and deethylatrazine were the only organic compounds detected at the four wells. Atrazine was detected in samples from two of the four wells (Hn-1917 and Li-0656) and deethylatrazine was detected in samples from all four wells. Concentrations of atrazine and deethylatrazine detections were all less than 0.05 μg/L (Ging and Mobley, 2021). Atrazine is a triazine herbicide commonly used to control broadleaf and grassy weeds in a variety of crops such as corn and sorghum (EPA, 2021) and has a drinking-water standard of 3.0 µg/L (EPA, 2009). Deethylatrazine is a pesticide degradate of atrazine (EPA, 2006).

During the 2012–13 sampling period, samples from the same six wells were analyzed for the same organic compounds at NWQL and OGRL. Three of the four wells where organic compounds were detected in the samples collected during the 2019–20 sampling period were the same wells where organic compounds were detected in the sample collected during the 2012–13 sampling period. Atrazine, deethylatrazine, and tebuthiuron were detected during the 2012–13 sampling period, and concentrations were all less than 0.03 μg/L, which is slightly less than the detected concentrations during the 2019–20 sampling period. Tebuthiuron was not detected during the 2019–20 sampling period (Ging and Mobley, 2021).