EDITORIAL OFFICE: Environmental & Engineering Geoscience journal, Kent State University, Kent, OH 44242, U.S.A., ashakoor@kent.edu. CLAIMS: Claims for damaged or not received issues will be honored for 6 months from date of publication. AEG members should contact AEG, 3053 Nationwide Parkway, Brunswick, OH 44212. Phone: 844-331-7867. GSA members who are not members of AEG should contact the GSA Member Service center. All claims must be submitted in writing. POSTMASTER: Send address changes to AEG, 3053 Nationwide Parkway, Brunswick, OH 44212. Phone: 844-331-7867. Include both old and new addresses, with ZIP code. Canada agreement number PM40063731. Return undeliverable Canadian addresses to Station A P.O. Box 54, Windsor, ON N9A 6J5 Email: returnsil@imexpb.com. DISCLAIMER NOTICE: Authors alone are responsible for views expressed in articles. Advertisers and their agencies are solely responsible for the content of all advertisements printed and also assume responsibility for any claims arising therefrom against the publisher. AEG and Environmental & Engineering Geoscience reserve the right to reject any advertising copy. SUBSCRIPTIONS: Member subscriptions: AEG members automatically receive digital access to the journal as part of their AEG membership dues. Members may order print subscriptions for $75 per year. GSA members who are not members of AEG may order for $60 per year on their annual GSA dues statement or by contacting GSA. Nonmember subscriptions are $310 and may be ordered from the subscription department of either organization. A postage differential of $10 may apply to nonmember subscribers outside the United States, Canada, and Pan America. Contact AEG at 844-331-7867; contact GSA Subscription Services, Geological Society of America, P.O. Box 9140, Boulder, CO 80301. Single copies are $75.00 each. Requests for single copies should be sent to AEG, 3053 Nationwide Parkway, Brunswick, OH 44212. © 2023 by the Association of Environmental and Engineering Geologists All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or by any information storage and retrieval system, without permission in writing from AEG.
EDITORS ABDUL SHAKOOR Kent State University Kent, OH 44242 ashakoor@kent.edu
ERIC PETERSON Department of Geography, Geology, and the Environment Illinois State University Normal, IL 61790 309-438-5669 ewpeter@ilstu.edu
KAREN E. SMITH, Editorial Assistant, kesmith6@kent.edu
OOMMEN, THOMAS Board Chair, Michigan Technological University SASOWSKY, IRA D. University of Akron
ASSOCIATE EDITORS ACKERMAN, FRANCES Ramboll Americas Engineering Solutions, Inc. BASTOLA, HRIDAYA Lehigh University BEGLUND, JAMES Montana Bureau of Mines and Geology BRUCKNO, BRIAN Virginia Department of Transportation CLAGUE, JOHN Simon Fraser University, Canada DEE, SETH University of Nevada, Reno FRYAR, ALAN University of Kentucky GARDNER, GEORGE Massachusetts Department of Environmental Protection GHOSH, SUMAN California Department of Conservation
HAUSER, ERNEST Wright State University KEATON, JEFF WSP USA MAY, DAVID USACE-ERDC-CHL POPE, ISAAC Book Review Editor SANTI, PAUL Colorado School of Mines SCHUSTER, BOB SHLEMON, ROY R.J. Shlemon & Associates, Inc. STOCK, GREG National Park Service SWANSON, SUSAN (SUE) Beloit College ULUSAY, RESAT Hacettepe University, Turkey WEST, TERRY Purdue University
Environmental & Engineering Geoscience AUGUST 2023
VOLUME XXIX, NUMBER 3
Special Issue on Karst: Guest Editors – Cory Blackeagle, Russell Harmon, and Robert Denton
Submitting a Manuscript Environmental & Engineering Geoscience (E&EG), is a quarterly journal devoted to the publication of original papers that are of potential interest to hydrogeologists, environmental and engineering geologists, and geological engineers working in site selection, feasibility studies, investigations, design or construction of civil engineering projects or in waste management, groundwater, and related environmental fields. All papers are peer reviewed. The editors invite contributions concerning all aspects of environmental and engineering geology and related disciplines. Recent abstracts can be viewed under “Archive” at the web site, “http://eeg.geoscienceworld.org”. Articles that report on research, case histories and new methods, and book reviews are welcome. Discussion papers, which are critiques of printed articles and are technical in nature, may be published with replies from the original author(s). Discussion papers and replies should be concise. To submit a manuscript go to https://www.editorialmanager.com/EEG/ default.aspx. If you have not used the system before, follow the link at the bottom of the page that says New users should register for an account. Choose your own login and password. Further instructions will be available upon logging into the system. Upon submission, manuscripts must meet, exactly, the criteria specified in the revised Style Guide found at https://www.aegweb.org/e-eg-supplements. Manuscripts of fewer than 10 pages may be published as Technical Notes. The new optional feature of Open Access is available upon request for $750 per article. For further information, you may contact Dr. Abdul Shakoor at the editorial office.
Cover photo To understand carbon dioxide (CO2) fluxes in karst soils, three soil profiles were instrumented with CO2, temperature, and water content sensors near the entrance to James Cave, Virginia. Photo credit: Taryn Thompson. See article on page 217.
Volume XXIX, Number 3, August 2023
THIS PUBLICATION IS PRINTED ON ACID-FREE PAPER
ADVISORY BOARD WATTS, CHESTER “SKIP” F. Radford University HASAN, SYED University of Missouri, Kansas City NANDI, ARPITA East Tennessee State University
ENVIRONMENTAL & ENGINEERING GEOSCIENCE
Environmental & Engineering Geoscience (ISSN 1078-7275) is published quarterly by the Association of Environmental & Engineering Geologists (AEG) and the Geological Society of America (GSA). Periodicals postage paid at AEG, 3053 Nationwide Parkway, Brunswick, OH 44212 and additional mailing offices.
THE JOINT PUBLICATION OF THE ASSOCIATION OF ENVIRONMENTAL AND ENGINEERING GEOLOGISTS AND THE GEOLOGICAL SOCIETY OF AMERICA SERVING PROFESSIONALS IN ENGINEERING GEOLOGY, ENVIRONMENTAL GEOLOGY, AND HYDROGEOLOGY
Environmental & Engineering Geoscience Volume 29, Number 3, August 2023 Table of Contents 155
Foreword Cory Blackeagle, Russell Harmon, and Robert Denton
157
Adherence of Polystyrene Microspheres on Cave Sediment: Implications for Organic Contaminants and Microplastics in Karst Systems Jill L. Riddell, Dorothy J. Vesper, and Louis M. McDonald
169
Potential of Cave Sediments as a Proxy for Tropical Cyclone and Storm Activity Jason Polk and Philip van Beynen
183
Recommended Planning and Response for Hazardous Material Releases in Karst Terrains Geary M. Schindel, Rudolph Rosen, and Graham M. Schindel
191
Economic Exclusion and Forgotten Floodplains on Karst Terrain Sarah A. Burgess and Lee J. Florea
203
Hydrochemical Delineation of Spring Recharge in an Urbanized Karst Basin, Central Kentucky (USA) Alan E. Fryar, Benjamin J. Currens, and Cristopher S. Alvarez Villa
217
Assessing Landscape and Seasonal Controls on Soil CO2 Fluxes in a Karst Sinkhole Taryn K. Thompson, Daniel L. McLaughlin, Madeline E. Schreiber, and Ryan D. Stewart
Foreword for E&EG Karst Special Edition CORY BLACKEAGLE RUSSELL HARMON ROBERT DENTON
The term karst, in use since the mid-18th century to describe the limestone plateau region adjacent to the Bay of Trieste in Slovenia and in Italy, refers to a collection of features of a natural landscape formed by dissolution of rock, most commonly carbonates, evaporites, and, less commonly, marbles and quartzites, by naturally occurring acidic fluids. Enclosed depressions, sinking streams, springs, and caves are common features of all karst landscapes, but weathering under different climatic regimes over varying geological time scales has produced a broad variety of karst landscapes that range from youthful mountainous situations characterized by steep topography and deep canyons to mature terrains comprising broad undulating plains. The term pseudokarst is applied to cave features formed via other processes (e.g., volcanic lava tubes, sea caves, talus caves, etc.). The papers in this issue focus on dissolutional karst in carbonate rocks. Twenty-five percent of the global land surface is underlain by karst. This highly dynamic landscape serves as a fragile foundation for both urban and rural populations. The tapping of karst waters for drinking water has been important in the historical and economic development of many karst regions, and today, large karst aquifers provide an important source of water worldwide. Although karst caves have been utilized since the earliest days of humankind and have long captured the popular imagination, most people in today’s modern world are unaware of karst systems and their importance. One unique characteristic of karst landscapes is that the movement of surface water into the subsurface can be very rapid and occur without the natural filtration provided by the soil and rock cover present in most other landscapes. Consequently, whatever is in the water and on the ground over which the water flows will be transported, unchanged, into the subsurface. This makes karst aquifers highly susceptible to both chemical and biological pollution. Thus, according to the U.S. Environmental Protection Agency, the unique characteristics of karst terrane and the aquifers it contains make karst aquifers the groundwater type most vulnerable to hazardous contaminants and pollution. About one-quarter of the world’s population, i.e., some 2 billion people, depends largely or entirely on groundwater obtained from karst. The world’s largest springs and most productive groundwater supplies are karstic, yet water resources in karst areas are the most easily polluted but only infrequently protected. Water
quantity and quality in karst waters can change rapidly and dramatically in time and distance. Karst aquifers are extremely heterogeneous complex systems, both hydrologically and hydrochemically. Pointsource recharge, via sinkholes, fractures, and fissures, is common, and the integrated triple porosity character is a unique feature of karst aquifers that makes them difficult to model. The current understanding of flow and transport in karst systems is incomplete. Nevertheless, a karst hydrogeology toolbox that comprises geological, geophysical, speleological methods, hydrologic and hydraulic techniques, and the use of hydrochemical, isotopic, and artificial tracers has been developed that can be used in the context of a particular conceptual model. Water storage in karst aquifers can vary from completely absent to highly efficient, with time frames of thousands of years and with its character and flow mechanisms varying from one karst site to another because of the physical peculiarities of individual karst systems. The complexity and variability make generalization difficult and potentially dangerous. In karst terranes, contaminants can be transmitted miles from their source and re-emerge, without dilution, within minutes or hours, in locations completely unanticipated. Among karst scientists, the catch phrase of “what goes down must come up” has long been used to summarize how water - and anything that water carries - flows through karst aquifers. Perhaps even more importantly, though, this simple phrase serves as a strong warning about how easily contaminants may appear in and pollute karst wells and springs. Karst terranes are particularly sensitive environments to a variety of both natural and anthropogenic hazards and their dynamic natures and unique characters can present hazards to both human life and built infrastructure. Flooding in karst terrains can be frequent, unpredictable, extensive, and dramatic. Land surface subsidence and collapse can occur with little or no warning, endangering people and causing costly damage to infrastructure. The surface-subsurface linkage is intimate, complicated, and extremely variable, spatially and temporally. Also, many rare and endangered flora and fauna exist solely in karst environs. Destabilization of this linkage can lead to ecosystem collapse. The study of karst terranes necessarily involves a variety of academic disciplines that include the geosciences and its many subdisciplines (geology, geochemistry, geophysics, structural geology, geomorphology, hydrology,
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 155–156
155
BlackEagle, Harmon, and Denton
speleology, paleoclimatology, and even planetary studies), engineering, archaeology, atmospheric science, and urban planning. Comprehensive karst studies can require the assistance of cave explorers and mappers, cave divers, mathematicians, and computer programmers and modelers. In all cases, practitioners in each discipline bring with them their own experiences, perspectives, insights, tools, and scales of reference. Therefore, most karst research and practical investigations require a multi- and interdisciplinary approach. Unfortunately, the vast majority of geoscientists and engineers are rarely, if at all, exposed to the complexity of karst terrane in their education, receiving, perhaps, an individual class in one course during their entire academic training. When presented with a karst terrane in professional practice, this lack of exposure and knowledge can be particularly problematic. This special edition reflects the multi-disciplinary nature of karst terrane research with papers focused on geochemistry, hydrogeology, sedimentology, paleoclimatology, flood risk mapping, hydrochemistry, CO2 fluxes, soil properties, and HAZMAT protocols. Riddell, Vesper, and McDonald used polystyrene microspheres as a proxy for bacteria transport in karst groundwater. Others have used microplastics in this way, but the results of these studies have been hampered by a loss of microplastics in the course of tracing studies (i.e., more are released than are recovered). A key finding is that pH and mineral content of the sediment have the greatest impact on adherence of the microspheres to the sediment, which will slow transport and increase storage through the karst system. This illustrates the complexity of understanding and predicting the transport of organic and/or microplastic contaminants in karst systems. Caves act as repositories of clastic sediment both from normal and extraordinary processes and the caves protect the sediment from reworking and other forms of disturbance. This property is becoming increasingly recognized and valued for paleoclimate reconstructions. Polk and van Beynen present a case for the use of cave sediments as a proxy to study paleo storm activity by
156
examining the litho- and biostratigraphy of sediment records from two caves in Florida. Schindel, Rosen, and Schindel recommend that prevention is the best course to prevent contamination of karst aquifers, particularly those that serve as water supplies. They provide a set of best management practices for resource managers and first responders to accomplish that. As a result of the flashy and unpredictable manner that flooding occurs in karst terrane and the lack of knowledge of karst terrane, areas that should be included on maps for the U.S. National Flood Insurance Program are frequently excluded. Burgess and Florea utilize a combination of geographic information systems (GIS), historical flood documentation, stream gauging, and hydrography to provide a substantially improved assessment of flood risk in an area excluded from coverage by the U.S. National Flood Insurance Program and establish a methodology for so doing. Fryar, Currens, and Villa utilize the geochemical and physical properties of spring water, along with hydrographs, to differentiate non-point and point input recharge into a karst groundwater basin. The sinkhole that provided point recharge to the groundwater basin underwent restoration prior to the study, and they concluded that the restoration improved stormwater management. It is important to understand the diffusion of CO2 gas through soils in karst terranes because it controls the formation of carbonic acid within the critical zone as well as how CO2 moves through the carbon cycle. Thompson and colleagues investigated the possible existence of a zero-flux CO2 plane in the shallow subsurface. They observed that soil in different locations within a sinkhole displayed seasonally distinct CO2 concentrations and proposed that, as a result of convective venting during the cool season, the primary CO2 sink in the system was the underlying cave connected to the sinkhole. We are very grateful to the authors who provided these papers, and hope that readers will find them to be interesting and informative.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 155–156
Adherence of Polystyrene Microspheres on Cave Sediment: Implications for Organic Contaminants and Microplastics in Karst Systems JILL L. RIDDELL* Chatham University Falk School of Sustainability and Environment, Gibsonia, PA, USA, 15044
DOROTHY J. VESPER Department of Geology and Geography, West Virginia University, Morgantown, WV 26505
LOUIS M. MCDONALD Division of Plant and Soil Sciences, West Virginia University, Morgantown, WV 26505
Key Terms: Microspheres, Cave Sediments, Sorption, Adherence, Microplastics, Karst
sediment interactions with potential organic or MP contaminants in karst systems.
ABSTRACT
INTRODUCTION
Interactions of karst aquifer sediments with organic contaminants or microplastics (MPs) have received relatively little attention even though the susceptibility of karst aquifers to contamination and their ability to store and transport sediment are well documented. Studies using polystyrene microspheres as surrogate tracers for bacteria transport in karst systems have generally observed low recovery of microspheres and attributed this to microsphere adsorption onto aquifer sediments. In addition to being used as surrogate tracers for bacteria, microspheres have the potential to be used as surrogate material for organic contaminants and MPs. Using cave sediments as a proxy for karst aquifer sediments, the adherence of two types of microspheres (carboxylated and non-functionalized) was measured in three different types of solutions: deionized water (DI), a calcium carbonate solution, and a karst spring water. Both types of microspheres adhered to the sediments; the most influential factor in adherence was solution type not microsphere type. Average adherence ranged from 51 to 94 percent with average adsorption coefficients (KD) ranging 11.8–442. Average estimated organic carbon–water partition coefficients (KOC) and retardation factors (RF) ranged from 1.64 3 103 to 6.13 3 104 and from 6.20 3 101 to 2.29 3 103, respectively. KD, KOC, and RF were an order of magnitude higher in the karst water than in DI or CaCO3 solution. The results illustrate the importance of
The contamination of karst waters by microplastics (MPs) and organic chemicals has received increased attention in recent decades (Ewers et al., 1991; Crawford and Ulmer, 1994; Vesper, 2002; Vesper et al., 2003; Loop and White, 2005; Padilla et al., 2011; Ewers et al., 2012; Ghasemizadeh et al., 2015; White et al., 2018; Panno et al., 2019; and Balestra and Bellopede, 2021), but the role of karst aquifer sediments and their interactions with these contaminants has received comparatively less attention. MPs are chemically organic solid particles ,5 mm in diameter that are derived from the production and breakdown of primary plastic materials (Prata et al., 2019; Zhou et al., 2019; and Corami et al., 2020). MPs have become ubiquitous in the environment (Petersen and Hubbart, 2021) and present a challenging pollution problem due to their ability to be the sorbate for organic chemicals and metals and carry those contaminants into other systems (Petersen and Hubbart, 2021). Recent studies in a show cave in Italy (Balestra and Bellopede, 2021) and in the karst region of Illinois (Panno et al., 2019) detected measurable amounts of MPs in sediments (Balestra and Bellopede, 2021) and groundwater (Panno et al., 2019). Still, the surface chemical interactions between MPs and karst sediments as well as the larger role of karst sediments in contaminant fate and transport remain relatively unquantified. Caves provide an accessible window into the larger karst aquifer system. Clastic cave sediments are a useful and abundant resource in understanding the role of karst aquifer sediments in contaminant fate and transport. Karst
*Corresponding author email: j.riddell@chatham.edu
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
157
Riddell, Vesper, and McDonald
aquifers can transport sediments during high-velocity transport events during which sediments can enter the system (Doehring and Vierbuchen, 1971; Herman et al., 2008), while existing sediments can be re-mobilized and transported within the system or to outlet points, such as springs (Mahler, 1999; Herman et al., 2008). Sediments can be a contaminant or act as a carrier of adsorbed contaminants through karst systems (McCarthy and Zachara, 1989; Mahler et al., 1999; Frimmel et al., 2007; Mahler et al., 2007; Goeppert and Hoetzl, 2009; and Goeppert and Goldscheider, 2019). Particle tracer experiments (Kass, 1992) commonly use natural particles (e.g., sediments) or engineered particles (e.g., microspheres) to evaluate particle movement through an aquifer. Polystyrene microspheres are manufactured particles ranging in size from ,1 lm to 1,000 lm, and they can be made with a range of surface charges and fluorescent tags, which influence their chemical behavior. Due to their similar dimensions to bacteria, polystyrene microspheres have been used in tracer experiments to understand microbe transport in karst aquifers (Harvey et al., 1989; Lindqvist and Bengtsson, 1995; Auckenthaler et al., 2002; Goldscheider et al., 2003; Goeppert and Goldscheider, 2008, 2011; Harvey et al., 2008; Sinreich et al., 2009; Ward et al., 2016; and Bandy et al., 2020). Recovery of microspheres in these experiments ranged from 0 to 57 percent with a median value of 5 percent. Low recovery was often attributed to adsorption of the microspheres onto sediments or mineral surfaces. Some research has attempted to describe and quantify microsphere adsorption as it relates to bacterial adsorption (Lindqvist and Bengtsson, 1995; Sinreich et al., 2009), but few studies have quantified microsphere adsorption onto karst sediments and directly compared those results to field experiments. Furthermore, the potential for microspheres to act as surrogates for specific characteristics of organic contaminants or MPs has not yet been examined. The purpose of this study was to: (i) conduct batch adherence experiments to determine if polystyrene microspheres adhere to clastic cave sediments and quantify any adherence; (ii) conduct a preliminary investigation into the chemical drivers of microsphere adherence; and (iii) use the results of the adherence experiments to estimate the adsorption coefficients (KD) and organic carbon–water partitioning coefficients (KOC). The term adsorption describes the adhesion of a sorbate (molecules or ions) in a dissolved state onto a solid surface, which is called the sorbent (Hillel, 2008; Owens and Rutledge, 2005). In the case of these experiments, the term adherence, rather than adsorption, is used since both the sorbate (microspheres) and sorbent (sediment) were in the solid phase. The sorbate (1 lm microspheres, in this case) 158
was also much larger relative to other common sorbates (dissolved metals, aqueous phase molecules, etc., which are usually 0.0001 0.001 lm). METHODS Sediment Preparation and Analysis A single, composite clastic cave sediment was used as the sorbent for the adherence experiments. The sediment was collected in bulk from McClung-Zenith Cave (also called Dropping Lick Spring or Dropping Lick Cave) in Monroe County, WV (WVaSS, 2019). This cave is formed in an Ordovician dolostone and limestone unit and has an active cave stream (WVaSS, 2019). The sediments were all above the water level at the time of collection. The sediment was air dried in a dark room for 72 hours and then lightly ground and sieved to , 2 mm (sand size and smaller; Owens and Rutledge, 2005). Particles in this size range have high surface area and are responsible for surface chemistry interactions within a sediment profile (Hillel, 2008). This composite sediment was used in all the experiments. Sand-sized particles made up 35–61 percent and siltsized particles made up 39–65 percent of the ,2 mm fraction; no clay-sized particles were detected (Riddell, 2022). Samples were classified as very fine sand, very coarse silt, or very coarse silty very fine sand. Quantitative X-ray diffraction (qXRD) identified quartz (67 percent), dolomite (5.9 percent), K-feldspar (5.6 percent), chlorite (4.3 percent), and amorphous material (17 percent) as the most abundant components in the sediment (Riddell, 2022). Total carbon (TC) and total inorganic carbon (TIC) for the ,2 mm fraction were measured on a Carlo Erba NA1500 CNHS elemental analyzer at the University of Florida Stable Isotope Mass Spectroscopy Laboratory in Gainesville, FL. TIC was measured by acidifying the sediment and quantifying the degassed CO2 using an UIC 5017 CO2 coulometer. Total organic carbon (TOC) was determined as the difference between TC and TIC. TC, TIC, and TOC are reported as dry weight percent by sample. Microsphere Selection Fluoresbrite® polystyrene 1.0 lm yellow-green (YG) fluorescent (excitation ¼ 441 nm, emission ¼ 486 nm) carboxylated microspheres (CMS) and non-functionalized microspheres (NFMS) were selected for the experiments (Polysciences, Inc., Warrington, PA; respective item numbers 15702-10 and 17154-10). Polystyrene is a synthetic aromatic hydrocarbon polymer consisting of a hydrocarbon chain attached to a benzene ring (Figure 1a). Polystyrene is a component in many plastic and foam products and was chosen as the base material for these experiments due to its use in past tracer experiments and relative
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
Polystyrene Microspheres on Cave Sediment
Figure 2. Flow chart of experimental design and solution preparation. Figure 1. (a) Chemical structure of polystyrene (C8H8, International Union of Pure and Applied Chemistry ID poly[1-phenylethene-1,2diyl]) microspheres and scanning electron microscope (SEM) image of NFMS. (b) Chemical structure of COOH group and SEM image of CMS.
chemical inertness in water. The CMS are modified with a carboxyl (COOH) functional group attached to the hydrocarbon chain (Figure 1b), which alters the chemical properties and behavior of the CMS in solutions. Both types of microspheres are polar compounds and can participate in hydrogen bonding, which influences adherence. COOH groups are proton donors with an acid dissociation constant (pKa) around pH ¼ 5 (Wade, 2006), and they can participate in ionic interactions. Some evidence shows that the pKa of a carboxylic acid group can increase by up to 0.43 units with each addition of a CH2 group onto the chain (Vysotsky et al., 2020). Thus, in the case of CMS, the pKa may be higher than 5. When the pH pKa, the COOH group deprotonates, and a negatively charged surface (COO ) is created. This negatively charged surface contributes a degree of hydrophilicity to the molecule and may result in less interaction between the CMS and other particles or compounds in solution. However, the negatively charged surface on the CMS may also interact with any positively charged surfaces in the sediment or solution. The NFMS, while labeled neutral or non-functionalized due to the absence of an attached functional group, also have a slightly negative surface charge (Lundqvist et al., 2008), which may promote their interactions with dissolved or solid cationic species in the sediment solution. Positively and negatively charged polystyrene microspheres have also been shown to attach to soils depending on the zeta potential of the soils (Wang et al., 2022). Furthermore, the hydrophobicity of the benzene ring will drive both molecules out of solution and encourage adherence. Experimental Design and Data Analysis To quantify microsphere adherence to the sediments, batch isotherm experiments were conducted according to the procedure published by Roy et al. (1991), which described batch experimental techniques for testing the
sorption capacity of different soils with ionic and organic compounds. The experimental procedure is illustrated in Figure 2. Both types of microspheres were purchased in a 2.5 percent aqueous suspension in 10 mL quantities at a concentration of 4.55 3 1010 particles/mL (Fluoresbrite Carboxy YG 1 lm microspheres [MSDS 15702; https://www .polysciences.com/default/awfile/index/attach/file/15702.pdf] and Fluoresbrite Plain YG 1 lm microspheres [MSDS 17154, https://www.polysciences.com/default/awfile/index/ attach/file/17154.pdf/], Polysciences, Inc., Warrington, PA.). Because of these high concentrations in small volumes, serial dilution of 5,0003 was necessary to create a stock solution of microspheres that could be accurately counted during each stage of the experiments. Microsphere concentrations are reported as spheres per milliliter (sph/mL). For both CMS and NFMS, dilutions were made in the following solutions: (i) organic-free, deionized water (DI), (ii) a 25 mg/L CaCO3 solution, and (iii) an unfiltered representative karst water, which provided a total of six experimental types (two microsphere types and three solution types). The karst water was collected from a tufa spring in southwestern Pennsylvania and analyzed for dissolved calcium and magnesium according to Environmental Protection Agency (EPA) Methods 300.0 (USEPA, 1993) at the West Virginia University National Resource Center for Coal and Energy. Alkalinity and ionic strength (IS) of the DI, CaCO3 solution, and karst water were estimated using U.S. Geological Survey geochemical modeling software PHREEQC (Parkhurst and Appelo, 2013) assuming a charge balance error of zero. Estimations were made using software due to the low volume of the solutions available for experiments and analyses. Up to 14 experimental solutions with varying concentrations of microspheres were used. Initial concentrations of these solutions ranged from 3.41 3 106 sph/mL to 5.21 3 102 sph/mL. These microsphere solutions were mixed with the representative sediment to measure the adherence of each type of microsphere onto the sediment in each experimental solution. The chemistry of the different water types is outlined in Table 1.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
159
Riddell, Vesper, and McDonald Table 1. Chemistry of water types used in microsphere dilutions. Parameter pH Mg (mg/L) Ca (mg/L) Ionic strength (mol/L) Alkalinity (meq/L)
DI Water
25 mg/L CaCO3
Karst Water
5.62 0 0 — —
7.27 0 25 5.38 3 10 4 7.52 3 10 5
8.29 32.9 74.0 1.03 3 10 3 0.99
All solutions were in equilibrium with atmospheric conditions.
The sediment-to-solution ratio and equilibration time were determined via the methods described in Roy et al. (1991) for ionic solutes. (CMS can participate in ionic bonding due to the COOH group, so this method was selected for consistency across the experiments.) A ratio of 1:20 ( 1.25 g sediment to 25 mL solution) and an equilibration time of 4 hours were chosen based on preliminary optimization tests. These parameters resulted in adhered microspheres consistent with previously reported field experiments and allowed for comparison of different microsphere behavior under the same conditions. The initial solutions were mixed in amber glass vials under the darkest possible laboratory lighting to minimize degradation of microspheres under fluorescent light. The solutions were then rotated on a Glas-Col Rugged Rotator (Terre Haute, IN) rotary shaker at 70 rotations per minute (RPM) for 4 hours. Subsequently, the solutions were centrifuged at 1,000 RPM (88.5 g-force) on a Beckman Coulter Allegra X-30 series centrifuge to separate the supernatant from the sediment mixtures. The supernatant (equilibrium solution) was then pipetted into a clean amber glass container for storage until analysis. The pH of the initial and equilibrium solutions was measured on a Hanna Instruments (Smithfield, RI) benchtop pH/oxidation-reduction potential (ORP)/ion-selective electrode (ISE) meter. The initial and equilibrium solutions were analyzed on a BD LSR Fortessa Analyzer flow cytometer at the West Virginia University Flow Cytometry and Single Cell Core Facility. This instrument counts cells and particles in a fluid stream. The instrument uses fluorescence-activated cell sorting to sort particles of interest (in this case, 1 lm YG microspheres) as the sample flows past an excitation source (laser). The emitted fluorescence is detected by a forward scatter (FCS) diode, and the resulting particle information is reported. All particle events in a sample are counted. This procedure provided an exact count of microspheres in 50 lL of solution, from which the equilibrium and initial concentrations of sph/mL were determined. Data Evaluation The microsphere counts from the flow cytometer were used to calculate the adhered concentration of microspheres 160
(sph/g) relative to the mass of sediment used according to Equation 1: sph sph sph Ad conc ¼ init equil g mL mL 3
volume solutionðmLÞ ; mass sediment ðgÞ
(1)
where ad conc is the adhered concentration, init is the initial solution concentration, and equil is the equilibrium solution concentration. The KD of the microsphere solutions was calculated as the slope of the line of best fit on the plot of adhered concentration versus equilibrium concentration. The fraction of organic carbon (fOC) of the sediment was calculated based on the TOC. From this, the KOC was calculated as: KOC ¼
KD fOC
(2)
Because sorption or adherence can occur via partitioning onto soil or sediment or into organic matter (Schwarzenbach et al., 2003), KD values are highly variable based on the organic carbon content of the sorbent. Normalization of KD values to the fOC by calculating KOC allows for a better comparison between the same sorbates with different sorbents. Although the conversion to KOC was not critical in this study because all experiments used the same sediments, it allows the data to be compared to similar studies of other sediments. The calculated KD and KOC values for the microsphere experiments were compared to the known KOC values of contaminants in karst aquifers (e.g., chlorinated solvents, volatile organic compounds, and plasticizers) to estimate the potential for the microspheres to act as a surrogate for these contaminants in future experiments. One-way and two-way analysis of variance (ANOVA) tests were performed in R (R Core Team, 2019; Bevans, 2022) for each experiment to determine if any statistically significant relationship existed between KD and microsphere type or solution type. The same ANOVA test was performed for percent removed microspheres (that is, the microspheres removed from solution), microsphere type, and solution type. RESULTS Water Chemistry and Sediment Analysis of the Active Fraction, ,2 mm The pH of the solutions ranged from 5.62 (DI) to 8.29 (karst water, Table 1). The karst water had IS an order of magnitude higher than the CaCO3 solution and alkalinity
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
Polystyrene Microspheres on Cave Sediment
Figure 3. Experimental results for (a) NFMS and (b) CMS in DI, 25 mg/L CaCO3, and karst water. Error bars represent the standard error of replicate concentrations.
four orders of magnitude higher than that of the CaCO3. IS and alkalinity were negligible in DI. In seven replicates of the composite sediment, concentrations of weight percent TC, TIC, and TOC ranged 1.07–1.18 percent, 0.16–0.60 percent, and 0.57–0.91 percent, respectively. The average TOC concentration was 0.72 percent, and this was used for the fOC value (0.0072) in subsequent calculations. These TOC concentrations are in the same range as those reported for unsaturated clastic cave sediments elsewhere (Bottrell, 1996; Panno et al., 2004; de Paula et al., 2016, 2020; Downey et al., 2023; and Riddell et al., 2023). Adherence of NFMS and CMS Four experiments in DI and three each in 25 mg/L CaCO3 and karst water for CMS and NFMS were averaged together to determine microsphere behavior in the three solution types. For NFMS and CMS in all solutions, linear correlation was observed between the adhered and equilibrium concentrations, indicating adherence did occur (Figure 3a and b). An R2 .0.9 was observed in four out of six experimental configurations (Table 2). For all experiments, linear adherence was observed likely due to the low concentration of microspheres required for the flow cytometer analysis. The percent removal of NFMS and CMS was highest in the karst water (Table 2). However, removal was considerable in all experimental types, ranging 57–94
percent for NFMS and 50–94 percent for CMS. These removal rates were consistent with the rates of unrecovered microspheres in tracer experiments (e.g., Harvey et al., 2008; Goeppert and Goldscheider, 2011; and Bandy et al., 2020). In NFMS and CMS, the lowest KD values were reported in the CaCO3 solution experiments, followed by DI experiments. The karst water experiments had the largest KD by an order of magnitude (Figure 3a and b). The KOC for NFMS and CMS showed a similar pattern to the KD values, where the karst water experiments had KOC values an order of magnitude higher than the DI and 25 mg/L CaCO3 experiments. Using a two-way ANOVA test (two microsphere types and three solution types), a statistically significant difference (a ¼ 0.05) in the mean KD values was found for solution type (F ¼ 14.6, p ¼ 0.0003) but not for microsphere type (F ¼ 0.002, p ¼ 0.96). A similar ANOVA result was found for mean percent removed microspheres, where F ¼ 0.91 and p ¼ 0.48 for microsphere type, and F ¼ 11.3 and p ¼ 0.0009 for solution type. A test for homoscedasticity revealed the data for KD and percent removed microspheres to be non-normal, so a pair-wise test using the Wilcoxon sum rank test was used to identify any statistical difference between the solution types (regardless of sphere type, since sphere type had p .0.05). This test revealed a significant difference between the karst water and CaCO3 solutions (for KD and percent removed, p ¼ 0.003 in both tests) and for karst water and DI solutions, p ¼ 0.003 (KD) and p ¼ 0.002 (percent removed). There was no significant
Table 2. KD, KOC, and linear fit results of adherence experiments. Non-Functionalized microspheres (NFMS) Solution DI 25 mg/L CaCO3 Karst water
KOC
KD 64.9 11.8 252
R 3
9.02 3 10 1.64 3 103 3.52 3 104
2
0.90 0.70 0.97
Carboxylated microspheres (CMS) % Removed 57.1 58.2 93.5
Solution DI 25 mg/L CaCO3 Karst water
KD 38.5 26.2 442
KOC 3
5.35 3 10 3.64 3 103 6.13 3 104
R2
% Removed
0.96 0.98 0.73
60.8 50.8 93.8
KOC values were calculated based on the average fOC of the collected sediments, 0.0072.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
161
Riddell, Vesper, and McDonald
Figure 4. (a) Distribution of KD values regardless of microsphere type. Only solution type had a significant effect on KD values (p ¼ 0.0003), but microsphere type did not (p ¼ 0.9633). (b) Distribution of average percent removal regardless of microsphere type.
difference between DI and CaCO3 solutions. Regardless of microsphere type, the distribution of KD values indicated that the DI and CaCO3 solutions had similar KD medians and ranges, whereas karst solutions had higher KD median and larger range (Figure 4a). However, the average percent removed distribution (regardless of microsphere type) indicated that the karst solution removed more microspheres than the DI and CaCO3 solutions (Figure 4b). The results indicated that both types of microspheres behaved similarly in all solutions as shown by the KD, KOC values, and the statistical analysis of the KD and percent removed values. KD values were more consistent in the DI and CaCO3 solutions, but percent removal was more consistent in the karst solutions. An increase by an order of magnitude in both KD and KOC in the karst water experiments indicated that a potentially higher pH or increasing mineral content was driving both types of microspheres out of solution. Although both microspheres behaved similarly in each solution, the potential mechanisms for adherence may have been different among microsphere type, and this warrants further exploration. DISCUSSION Roles of IS, pH, and CaCO3 in Adherence Although both NFMS and CMS performed comparably in each solution type, the chemistry of each solution type may have affected microsphere adherence. The adherence of both types of spheres was likely being controlled by the pH, IS, and alkalinity of the different solution types, the contribution of pH and IS (and associated alkalinity) from the sediment, and, for the CMS, the pKa of the COOH functional group. Both microsphere types have some degree of hydrophobicity contributed by the polystyrene structure, which results in microspheres being driven out of solution (increased adherence). The CMS are amphiphilic, having both the hydrophobic properties of the polystyrene structure and the hydrophilic properties of the deprotonated COOH group at 162
high pH values. As pH increases across the water types presented here, so do IS and alkalinity (HCO3 ). Ions in solution, like Ca2þ, Mg2þ, and carbonate species (which were present in the CaCO3 solution and karst water solution), may complex with the COO group of CMS at high IS, making the CMS less hydrophilic and thereby increasing adherence. In solutions with lower pH and IS, the behavior between NFMS and CMS was similar. However, the DI solution resulted in a higher average KD than did the CaCO3 solution, even though pH and IS were the lowest in the DI solutions. This could indicate that some threshold value of pH or IS is required to observe the increase in KD in both NFMS and CMS. The effects of DI water on mineral surfaces can be difficult to estimate due to the low IS of DI water. The microspheres may have contributed some IS to the solutions reported here. Regardless of these effects, the DI experiments had similar KD results as the CaCO3 solution experiments. While IS and alkalinity were not calculated or measured in every solution due to low sample volumes, pH was measured in each initial and equilibrium solution for each experiment. This pH data can illustrate how the increase in IS and alkalinity may have affected adherence of microspheres. In the DI water experiments, the average pH of the DI water was 5.62, and the addition of the sediment to the water increased the average pH to 8.04 (Figure 5a). After the addition of NFMS into DI to create the initial NFMS solutions, the solution pH decreased (Figure 5a). When sediment was added to the NFMS solutions and brought to equilibrium, the pH increased (Figure 5a). After the addition of CMS to DI to create the initial CMS solutions, the pH increased (Figure 5a). After the sediment was added to the CMS initial solutions, the pH increased again (Figure 5a). The pH only exceeded 8 in CMS equilibrium solutions with low concentrations. This indicates that some property of the sediment resulted in an increase in pH, and some property of the NFMS contributed to a decrease in pH, while the addition of CMS resulted in an increase in pH. The CMS had a larger pH range than the NFMS in both the initial and equilibrium solutions. A higher concentration of CMS was associated with lower pH. This was likely a result of the dissociation of the Hþ from the COOH, since the pH values were .5 (the pKa of COOH). Thus, more CMS in a solution would result in a higher concentration of Hþ (lower pH). The negatively charged COO surfaces could also be a factor in microsphere adherence due to interactions with the positively charged surface on the sediments. However, adherence of NFMS was also observed in DI solution and could have been due to other intermolecular forces such as hydrogen bonding or Van der Waals forces.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
Polystyrene Microspheres on Cave Sediment
Figure 6. Relative concentrations of COOH and COO at different pH values from the experimental parameters.
Figure 5. Effect of CaCO3 and pH on microsphere adherence. The solid blue line represents the pH of the water type, and the dashed brown line represents the pH of the water with sediment at equilibrium. The black vertical lines represent the pH range in each solution type across the microsphere concentrations.
In the CaCO3 solution experiments, the initial pH of the CaCO3 solution of 7.27 was higher than that of the initial DI solution, but the addition of the sediment decreased the average pH of the CaCO3 solution to 6.24 (Figure 5b) rather than raising it, as in the case of the DI experiments. The initial solutions of NFMS and CMS had lower average pH ranges than the equilibrium solutions (Figure 5b). The CMS solutions had a larger pH range than the NFMS solutions. NFMS and CMS exhibited the same pattern of pH changes as in the DI solution experiments (Figure 5b) but over a smaller range of pH values. For the karst water experiments, the average pH ranges of all microsphere solutions (initial and equilibrium) were between 7.57 and 8.45 (Figure 5c). The pH of the karst water was 8.29 initially and then 8.06 after the addition of the sediment. These ranges of pH were much smaller than those in the DI and CaCO3 solutions and were likely due to the pH buffering capacity of karst
waters. Some buffering was likely also responsible for the smaller range of pH changes in the CaCO3 solution experiments as well. As pH exceeds 5, the proportion of deprotonated COO increases relative to the neutral COOH (Figure 6). All the solutions reported here were generally above pH 5, but the increasing concentration of COO at higher pH values could have been a contributing factor to the increase in adherence of CMS spheres in the karst waters relative to the DI and CaCO3 solutions because it resulted in interactions between positively charged particles on the sediment and the COO group. This does not explain the same increase that was observed in NFMS solutions. While several intermolecular forces, buffering capacity, and mineral surfaces in the sediment were likely working in concert to drive microsphere adherence under the different experimental conditions, these results show the role of pH changes driven by COOH dissociation and CaCO3 buffering capacity in microsphere adherence. Comparison to Microsphere Field Tracer Experiments Microsphere tracer tests reported in the literature nearly always use functionalized microspheres, given their similarity to bacteria (Lindqvist and Bengtsson, 1995; Harvey et al., 2008; Flynn and Sinreich, 2010; Goeppert and Goldscheider, 2011; and Bandy et al., 2020). The percent of unrecovered microspheres in these studies can be compared to the percent of removed microspheres reported here. Bandy et al. (2020) used 1.0 lm YG CMS to trace karst aquifers in central Kentucky to better understand Escherichia coli transport. Bandy et al. (2020) reported 88.7 percent unrecovered microspheres, which is comparable to the average percent of removed microspheres in karst water for NFMS (94 percent) and CMS (also 94 percent) reported here. A study in the northern Alps (Goeppert and Goldscheider, 2011) used functionalized 1.0 lm YG and red microspheres to estimate the transport of fecal indicator bacteria through a conglomerate carbonate (Goeppert and Goldscheider, 2011). The percent unrecovered YG and
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
163
Riddell, Vesper, and McDonald Table 3. Estimated retardation of microspheres through an aquifer. Microsphere and Solution
Estimated RF
NFMS DI solutions NFMS 25 mg/L CaCO3 solutions NFMS karst water solutions CMS DI solutions CMS 25 mg/L CaCO3 solutions CMS karst water solutions
3.37 3 102 6.20 3 101 1.30 3 103 2.00 3 102 1.37 3 102 2.29 3 103
RF was calculated according to Eq. 3. KD was directly calculated from experimental parameters; sediment density (q) and effective porosity (he) were estimated from values reported in Andersen et al. (2005), Grabowski et al. (2011), Stringer et al. (2016), and Woessner and Poeter (2020).
red microspheres was 43 percent and 94.9 percent, respectively. These values are comparable to the percent removed microspheres for NFMS and CMS in all solution types reported here. A study in the Biscayne aquifer of Florida used various sizes of functionalized microspheres to estimate the transport of cysts of Cryptosporidium, a common microbial groundwater contaminant (Harvey et al., 2008). That study reported unrecovered amounts of microspheres at 94.2 percent, 96.1 percent, and 97.1 percent, which are, again, comparable to the percent removed microspheres for NFMS and CMS karst water solutions reported here. Goeppert and Goldscheider (2011) and Harvey et al. (2008) speculated that low recovery of microspheres was due to adsorption onto aquifer sediments or rock matrix. This is supported by the results of the experiments reported here. The retardation factor (RF) of the microspheres in the tracer studies was not directly reported, but approximate retardation factors of the NFMS and CMS can be estimated. RF is directly related to the fOC, bulk density, and effective porosity of the substrate through which the microspheres are traveling. Rf ¼ 1 þ
q 3 KD ge
(3)
For the RF (Eq. 3) estimates of the NFMS and CMS, the fOC was directly measured, and bulk density (q) and effective porosity (he) were estimated from values for sandy sediments reported in Andersen et al. (2005), Grabowski et al. (2011), Stringer et al. (2016), and Woessner and Poeter (2020). Estimated RF ranged 62–1304 for NFMS and 137–2290 for CMS, where the largest value for both was in the karst solutions (Table 3). The estimated RF values for both NFMS and CMS were an order of magnitude higher in the karst water experiments than those in the DI or CaCO3 solution experiments, indicating a strong possibility that microspheres can be stored in karst systems during tracing experiments. This 164
experimental factor should be considered when using microspheres as a tracer in karst settings. Microspheres as a Surrogate Tracer for Organic Contaminants and MPs While microspheres may be similar in size to bacteria, they are organic compounds and may be a more suitable surrogate for particulate or colloidal organic contaminants. The potential for organic compounds to adsorb to sediments and soils and be stored means that karst sediments can act as a sink for organic contaminants. Furthermore, threshold transport events can dislodge sediments in karst aquifers and transport them out of the system (Doehring and Vierbuchen, 1971; Herman et al., 2008). Any associated contamination can then be reintroduced to the surface and potential ecological receptors, making karst aquifer sediments a contamination source as well. Bandy et al. (2020) and Ward et al. (2016) both observed remobilization of microspheres during highflow storm events. In addition to increased flow and subsequent sediment resuspension, the remobilization of microspheres during storm events could result from changes in aquifer water chemistry as it mixes with low IS and low pH rainwater. These chemical changes could result in de-adherence of the microspheres from the sediment. It is also possible that during tracing experiments, microspheres are becoming physically entrapped (rather than chemically entrapped via adherence) in lowtransmissivity, small-aperture fractures. The results from these tracer experiments and the batch adherence experiments presented here highlight, again, the necessity of robust sediment and water chemical characterization when undertaking contaminant research in karst settings. The mobility of organic contaminants through a sediment or soil is measured by their KOC value. Documented organic contaminants in karst aquifers include trichloroethylene (TCE), tetrachloroethylene (PCE), and various phthalates including di(2-ethylhexyl)phthalate (DEHP), diethyl phthalate (DEP), and di-n-butyl phthalate (DBP) (Padilla et al., 2011; Ghasemizadeh et al., 2015). The KOC values for these contaminants range from 64.3 (TCE) to 510,000 (DEHP) as reported by the EPA (1996). The estimated KOC values for NFMS and CMS here ranged from 103 to 104 and were in the same range as KOC values for phthalates (102–105), which are a component in plastics. Polystyrene is a material used in the production of plastic and foam goods, so it is possible, given the KOC range of NFMS and CMS reported here, that these microspheres could be used as a surrogate material to understand the behavior of organic contaminants or MPs in different soils and sediments. Emerging research on MPs has documented their presence in karst settings (Panno et al., 2019; Balestra
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
Polystyrene Microspheres on Cave Sediment
and Bellopede, 2021). Balestra and Bellopede (2021) developed a method to extract MPs from cave sediments using a density separation technique that accounts for the variation in the density of different polymers. That study quantified different sizes and shapes of MPs in the sediments. The data presented in this current study show that microspheres readily adsorb to sediments, yet the adherence or adsorption of MPs has rarely, if ever, been reported for cave sediments. The density separation technique presented by Balestra and Bellopede (2021) may result in underestimation of MPs in cave sediments if adhered MPs are not completely removed from the sediments. Panno et al. (2019) reported a concentration of 15.2 particles/L in karst waters from aquifers in Illinois. Sediments were not considered in that study, nor was the surface chemistry of the plastic particles that were identified. The current study indicates that, depending on surface chemistry, microspheres (or MPs with similar surface chemistry as microspheres) could be driven out of solution and adhere to sediments. Reported MP concentrations that were measured only in water in karst systems are likely underestimations (up to .90 percent) of MP contamination in the entire system. Although MP contaminants are not uniform in shape and size like the engineered microspheres, their likelihood to adhere onto sediments should be considered when quantifying MP contamination in karst systems. LIMITATIONS The adherence of polystyrene microspheres and organic compounds on soils and sediments has been modeled with non-linear isotherms (Xia and Ball, 1999; Arthur, 2020; and Wang et al., 2022). In the experiments presented in this study, low concentrations of microspheres were necessary for the counting function of the flow cytometer, and the results represent the linear portion of the isotherm. The true behavior of the microspheres is likely non-linear, yet the results still demonstrate the ability of microspheres to adhere to karst sediments. It was beyond the scope of this study to analyze MP contamination in the karst water and collected sediment. However, MP contaminants are often fibrous in shape (Corami et al., 2020; Balestra and Bellopede, 2021; and Petersen and Hubbart, 2021), while the microspheres used in these experiments were spherical. The flow cytometer is designed to specifically count spherical objects of a certain size and fluorescence (in this case 1 lm and YG fluorescence); as a result, the likelihood of MP contamination interfering with the results is negligible. Since many types of MP contaminants exist in the environment (Prata et al., 2019; Zhou et al., 2019; Corami et al., 2020; Balestra and Bellopede, 2021; and Petersen and Hubbart, 2021), the microspheres discussed here cannot completely represent the total behavior of every type of MP. Further
chemical characterization of the karst water (e.g., dissolved inorganic carbon analysis; repetition of the experiments in filtered vs. unfiltered water) may provide further insight into microsphere interactions with karst sediments. CONCLUSIONS Microsphere tracer experiments in karst aquifers have been used to understand the transport of bacteria, given their similarity in sizes. However, the tracer studies report consistently low recovery of the microspheres, which is commonly attributed to adsorption of microspheres onto aquifer sediments or matrix. Here, the adherence of two types of microspheres onto a clastic cave sediment was measured in laboratory batch experiments under different experimental conditions. The KOC and RF values of NFMS and CMS were comparable to known contaminants in karst aquifers such as phthalates. The microspheres adhered to the sediments regardless of microsphere type, although solution type had a statistically significant effect on KD and percent removed microspheres. The highest KOC values were calculated in the karst water experiments, suggesting that this experimental parameter results in the most adherence of microspheres. The KOC values of the microspheres were in the same range as those of phthalates. Analysis of the pH of each solution indicated that pH and the mineral content of the sediment were primary drivers of microsphere adherence to these sediments. The potential for microspheres to be used to understand surface chemistry interactions between organic contaminants or MPs and different soils and sediments warrants further exploration based on the results reported here. Estimations of MP contamination in the environment should consider adhered MPs in sediment and soils in addition to MP concentrations in water. ACKNOWLEDGMENTS The authors thank: Dr. Kathleen Brundage, Lindsey Knight, and the West Virginia University Flow Cytometry and Single Cell Core facility for use and instruction on the flow cytometer; Dr. Ellen Herman of Bucknell University for use of the particle size analyzer; Dr. Jason Curtis of the University of Florida for the total carbon analysis; and Autum Downey for assistance in sediment collection, processing, and particle size analysis.
Disclaimer This work was supported by the National Institute of Environmental Health Sciences Superfund Research Program project PROTECT (Puerto Rico Testsite for Exploring Contamination Threats) PTE Federal Award Number 5P42ES017198-12 and the West Virginia University Ruby Distinguished Doctoral
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
165
Riddell, Vesper, and McDonald Fellowship. The flow cytometry work was supported by: Mary Babb Randolph Cancer Center (MBRCC) Center of Biomedical Research Excellence (CoBRE) Grant GM103488, Tumor Microenvironment (TME) CoBRE Grant GM121322, and Fortessa S10 Grant OD016165. The authors have no relevant competing financial or non-financial interests to disclose. All authors contributed to the study conception and design. Material preparation, data collection, and analysis were performed by Jill Riddell. The first draft of the manuscript was written by Jill Riddell, and all authors commented on previous versions of the manuscript. All authors read and approved the final manuscript.
DATA AVAILABILITY Data are publicly available in the GitHub repository for this project: https://github.com/jlriddell12/microsphere adherence REFERENCES ANDERSEN, T. J.; LUND-HANSEN, L. C.; PEJRUP, M.; JENSEN, K. T.; AND MOURITSEN, K. N., 2005, Biologically induced differences in erodibility and aggregation of subtidal and intertidal sediments: A possible cause for seasonal changes in sediment deposition: Journal of Marine Systems, Vol. 55, No. 3, pp. 123–238. https://doi.org/10 .1016/j.jmarsys.2004.09.004. ARTHUR, B. K., 2020, Suitability of Differently Coated Fluorescent Microspheres for Simulating Sedimentation, Sorption and Transport Of Helminth Eggs During Wastewater Treatment or Soil Infiltration: Ph.D. Dissertation, Department of Geosciences, Ruhr University, Bochum, Germany, 123 p. AUCKENTHALER, A.; RASO, G.; AND HUGGENBERGER, P., 2002, Particle transport in a karst aquifer: Natural and artificial tracer experiments with bacteria, bacteriophages and microspheres: Water Science Technology, Vol. 43, No. 3, pp. 131–139. https://doi.org/10 .2166/wst.2002.0072. BALESTRA, V. AND BELLOPEDE, R., 2021, Microplastic pollution in show cave sediments: First evidence and detection technique: Environmental Pollution, Vol. 292, pp. 1–9. https://doi.org/10.1016/j .envpol.2021.118261. BANDY, A. M.; COOK, K.; FRYAR, A. E.; AND ZHU, J., 2020, Differential transport of Escherichia coli isolates compared to abiotic tracers in a karst aquifer: Groundwater, Vol. 58, No. 1, pp. 70–78. https://doi .org/10.1111/gwat.12889. BEVANS, R., 2022, ANOVA in R: A complete step-by-step guide with examples: Scribbr. Electronic document, available at https://www .scribbr.com/statistics/anova-in-r/ BOTTRELL, S. H., 1996, Organic carbon concentration profiles in recent cave sediments: Records of agricultural pollution or diagenesis?: Environmental Pollution, Vol. 91, No. 3, pp. 325–332. https://doi .org/10.1016/0269-7491(95)00064-x. CORAMI, F.; ROSSO, B.; BRAVO, B.; GAMBARO, A.; AND BARBANTE, C., 2020, A novel method for purification, quantitative analysis and characterization of microplastic fibers using micro-FTIR: Chemosphere, Vol. 238, pp. 1–10. https://doi.org/10.1016/j.chemosphere.2019.124564. CRAWFORD, N. C. AND ULMER, C. S., 1994, Hydrogeologic investigations of contaminant movement in karst aquifers of a train derailment near Lewisburg, Tennessee: Environmental Geology, Vol. 21, pp. 41–52. DE PAULA, C. C. P.; BICHUETTE, M. E.; AND SELEGHIM, M. H. R., 2020, Nutrient availability in tropical caves influences the dynamics of
166
microbial biomass: Microbiology Open, Vol. 9, No. 7, pp. 1–3. https://doi.org/10.1002/mbo3.1044. DE PAULA, C. C. P.; MONTOYA, Q. V.; RODRIGUES, A.; BICHUETTE, M. E.; AND SELEGHIM, M. H. R., 2016, Terrestrial filamentous fungi from Gruta do Catão (São Desidério, Bahia, northeastern Brazil) show high levels of cellulose degradation: Journal of Cave and Karst Studies, Vol. 78, No. 3, pp. 208–217. https://doi.org/10.4311/ 2016mb0100. DOEHRING, D. O. AND VIERBUCHEN, R. C., 1971, Cave development during a catastrophic storm in the Great Valley of Virginia: Science, Vol. 174, No. 4016, pp. 1327–1329. https://doi.org/10.1126/ science.174.4016.1327. DOWNEY, A. R.; RIDDELL, J. L.; PADILLA, I. Y.; AND VESPER, D. J., 2023, Storage and distribution of organic carbon in cave sediments: Examples from two caves in the northern karst region of Puerto Rico: Environmental Earth Sciences, Vol. 82, No. 36, pp. 1–14. https://doi.org/10 .1007/s12665-022-10720-2. ENVIRONMENTAL PROTECTION AGENCY (EPA), 1996, Soil Screening Guidance Technical Background Document Part 5: Chemical Specific Parameters: U.S. Environmental Protection Agency, electronic document, available at https://www.epa.gov/superfund/ superfund-soil-screening-guidance#technical EWERS, R. O.; DUDA, A. J.; ESTES, E. K.; IDSTEIN, P. J.; AND JOHNSON, K. M., 1991, The transmission of light hydrocarbon contaminants in limestone (karst) aquifers. In Stanley, A. and Quinlan, J. (Editors), Third Conference on Hydrology, Ecology, Monitoring, and Management of Ground Water in Karst Terranes: Association of Ground Water Scientists and Engineers, Nashville, TN, pp. 287–306. EWERS, R. O.; WHITE, K. A.; AND FULLER, J. F., 2012, Contaminant plumes and pseudo plumes in karst aquifers: Carbonates and Evaporites, Vol. 27, pp. 153–159. https://doi.org/10.1007/s13146012-0099-0. FLYNN, R.M. AND SINREICH, M., 2010, Characterisation of virus transport and attenuation in epikarst using short pulse and prolonged injection multi-tracer testing: Water Research, Vol. 44, No. 4, pp. 1138–1149. https://doi.org/10.1016/j.watres.2009.11.032. FRIMMEL, F. H.; VON DER KAMMER, F.; AND FLEMMIN, H. C., 2007, Colloidal Transport in Porous Media: Springer, New York, NY, 291 p. GHASEMIZADEH, R.; YU, X.; BUTSCHER, C.; HELLWEGER, F.; PADILLA, I.; AND ALSHAWABKEH, A., 2015, Equivalent porous media (EPM) simulation of groundwater hydraulics and contaminant transport in karst aquifers: PLoS One, Vol. 10, No. 9, pp. 1–21. https://doi.org/ 10.1371/journal.pone.0138954. GOEPPERT, N. AND GOLDSCHEIDER, N., 2008, Solute and colloid transport in karst conduits under low- and high-flow conditions: Groundwater, Vol. 46, No. 1, pp. 61–68. https://doi.org/10.1111/j.1745-6584 .2007.00373.x. GOEPPERT, N. AND GOLDSCHEIDER, N., 2011, Transport and variability of fecal bacteria in carbonate conglomerate aquifers: Groundwater, Vol. 49, No. 1, pp. 77–84. https://doi.org/10.1111/j.1745-6584 .2010.00741.x. GOEPPERT, N. AND GOLDSCHEIDER, N., 2019, Improved understanding of particle transport in karst groundwater using natural sediments as tracers: Water Research, Vol. 166. https://doi.org/10.1016/j.watres .2019.115045. GOEPPERT, N. AND HOETZL, H., 2009, Precise method for continuous measurement of fluorescent microspheres during flow: Hydrogeology Journal, Vol. 18, No. 2, pp. 317–324. https://doi.org/10.1007/ s10040-009-0517-0. GOLDSCHEIDER, N.; HÖTZL, H.; KÄSS, W.; AND UFRECHT, W., 2003, Combined tracer tests in the karst aquifer of the artesian mineral springs of Stuttgart, Germany: Environmental Geology, Vol. 43, No. 8, pp. 922–929. https://doi.org/10.1007/s00254-002-0714-9. GRABOWSKI, R. C.; DROPPO, I. G.; AND WHARTON, G., 2011, Erodibility of cohesive sediment: The importance of sediment properties:
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
Polystyrene Microspheres on Cave Sediment Earth Science Review, Vol. 105, No. 3–4, pp. 101–120. https://doi .org/10.1016/J.Earscirev.2011.01.008. HARVEY, R. W.; GEORGE, L. H.; SMITH, R. L.; AND LEBLANC, D. R., 1989, Transport of microspheres and indigenous bacteria through a sandy aquifer: Results of natural- and forced-gradient tracer experiments: Environmental Science and Technology, Vol. 23, pp. 51–56. https://doi.org/10.1021/es00178a005. HARVEY, R. W.; METGE, D. W.; SHAPIRO, A. M.; RENKEN, R. A.; OSBORN, C. L.; RYAN, J. N.; CUNNINGHAM, K. J.; AND LANDKAMER, L., 2008, Pathogen and chemical transport in the karst limestone of the Biscayne aquifer: 3. Use of microspheres to estimate the transport of potential of Cryptosporidium parvum oocysts: Water Resources Research, Vol. 44, No. 12. https://doi.org/10.1029/2007WR006060. HERMAN, E. K.; TORAN, L.; AND WHITE, W. B., 2008, Threshold events in spring discharge: Evidence from sediment and continuous water level measurement: Journal of Hydrology, Vol. 351, No. 1, pp. 98– 106. https://doi.org/10.1016/j.jhydrol.2007.12.001. HILLEL, D., 2008, Soil in the Environment: Soil Physical Attributes: Elsevier, Burlington, MA, 307 p. KASS, W., 1992, Tracing Technique in Geohydrology, 1st ed.: CRC Press, Boca Raton, FL, 602 p. LOOP, C. M. AND WHITE, W. B., 2005, A conceptual model for DNAPL transport in karst ground water basins: Groundwater, Vol. 39, No. 1, pp. 119–127, https://doi.org/10.1111/j.1745-6584.2001.tb00357.x. LINDQVIST, R. AND BENGTSSON, G., 1995, Diffusion-limited and chemical-interaction-dependent sorption of soil bacteria and microspheres: Soil Biology and Biochemistry, Vol. 27, No. 7, pp. 941– 948. https://doi.org/10.1016/0038-0717(95)00004-X. LUNDQVIST, M.; STIGLER, J.; ELIA, G.; LYNCH, I.; CEDERVALL, T.; AND DAWSON, K. A., 2008, Nanoparticle size and surface properties determine the protein corona with possible implications for biological impacts: Applied Physical Sciences, Vol. 105, No. 38, pp. 14265–14270. https://doi.org/10.1073/pnas.0805135105. MAHLER, B. J.; LYNCH, L.; AND BENNETT, P. C., 1999, Mobile sediment in an urbanizing karst aquifer: Implications for contaminant transport: Environmental Geology, Vol. 39, No. 1, pp. 25–38. https://doi .org/10.1007/S002540050434. MAHLER, B. J.; VALDES, D.; MUSGROVE, M.; AND MASSEI, N., 2007, Nutrient migration in carbonate aquifers in response to storms: A comparison of two geologically contrasting aquifers. Geological Society of America Abstracts with Programs, Vol. 39, No. 6, pp. 515. MCCARTHY, J. F. AND ZACHARA, J. M., 1989, Subsurface transport of contaminants: Mobile colloids in the subsurface environment may alter the transport of contaminants: Environmental Science and Technology, Vol. 23, No. 5, pp. 496–502. OWENS, P. R. AND RUTLEDGE, E. M., 2005, Encyclopedia of Soils in the Environment: Elsevier, New York, NY, 520 p. PADILLA, I.; IRIZARRY, C.; AND STEELE, K., 2011, Historical contamination of groundwater resources in the north coast aquifers of Puerto Rico: Revista Dimensions, Vol. 3, pp. 7–12. PANNO, S. V.; CURRY, B. B.; WANG, H.; HACKLEY, K. C.; LIU, C. L.; LUNDSTROM, C.; AND ZHOU, J., 2004, Climate change in southern Illinois, USA, based on the age and d13C of organic matter in cave sediments: Quaternary Research, Vol. 61, No. 3, pp. 301–313. https://doi.org/10.1016/j.yqres.2004.01.003. PANNO, S. V.; KELLY, W. R.; SCOTT, J.; ZHENG, W.; MCNEISH, R. E.; HOLM, N.; HOELLEIN, T. J.; AND BARANSKI, E. L., 2019, Microplastic contamination in karst groundwater systems: Groundwater, Vol. 57, No. 2, pp. 189–196. https://doi.org/10.1111/gwat.12862. PARKHURST, D. L. AND APPELO C. A. J., 2013, Description of Input and Examples of PHREEQC Version 3—A Computer Program for Speciation, Batch-Reaction, One-Dimensional Transport, and Inverse Geochemical Calculation: U.S. Geological Survey Techniques and Methods 6-A43, 497 p., https://pubs.usgs.gov/tm/06/a43/.
PETERSEN, F. AND HUBBART, J. A., 2021, The occurrence and transport of microplastics: The state of the science: Science of the Total Environment, Vol. 758, pp. 1–12. https://doi.org/10.1016/j .scitotenv.2020.143936. PRATA, J. C.; REIS, V.; MATOS, J. T. V.; DA COSTA, J. P.; DUARTE, A. C.; AND ROCHA-SANTOS, T., 2019, A new approach for routine quantification of microplastics using Nile Red and automated software (MP-VAT): Science of the Total Environment, Vol. 690, pp. 1277– 1283. https://doi.org/10.1016/j.scitotenv.2019.07.060. R CORE TEAM, 2019, R: A Language and Environment for Statistical Computing: R Foundation for Statistical Computing, Vienna, Austria, electronic document, available at https://ww.R-project.org/ RIDDELL, J. L., 2022, Chemical Characterization of Clastic Cave Sediments and Insights into Particle Transport and Storage in Karst Aquifers: Ph.D. Dissertation, Department of Geology and Geography, West Virginia University, Morgantown, WV, 157 p. https:// www.doi.org/10.33915/etd.11436. RIDDELL, J. L.; DOWNEY, A. R.; VESPER, D. J.; AND PADILLA, I. Y., 2023, Total organic carbon concentrations in clastic cave sediments from Butler Cave, Virginia, USA: Implications for contaminant fate and transport: Environmental Earth Sciences, Vol. 82, No. 231, pp. 1–16. https://doi.org/10.1007/s12665-023-10893-4. ROY, W. R.; KRAPAC, I. G.; CHOU, S. F. J.; AND GRIFFIN, R. A., 1991, Technical Resource Document: Batch-Type Procedures for Estimating Soil Adsorption of Chemicals: U.S. Environmental Protection Agency, Washington, D.C., 116 p. SCHWARZENBACH, R. P.; GSCHWEND, P. M.; AND IMBODEN, D. M., 2003, Environmental Organic Chemistry, 2nd ed.: Wiley and Sons, Hoboken, NJ, 1328 p. SINREICH, M.; FLYNN, R.; AND ZOPFI, J., 2009, Use of particulate surrogates for assessing microbial mobility in subsurface ecosystems: Hydrogeology Journal, Vol. 17, No. 1, pp. 49–59. https://doi.org/ 10.1127/1863-9135/2010/0177-0081. STRINGER, C. E.; TRETTIN, C. G.; AND ZARNOCH, S. J., 2016, Soil properties of mangroves in contrasting geomorphic settings within the Zambezi River Delta, Mozambique: Wetlands and Ecology and Management, Vol. 24, No. 2, pp. 139–152. https://doi.org/10.1007/ s11273-015-9478-3. USEPA, 1993, Method 300.0: Determination of Inorganic Anions by Ion Chromatography. U.S. Environmental Protection Agency, Cincinnatti, OH. VESPER, D. J., 2002, Transport and Storage of Trace Metals in a Karst Aquifer: An Example from Fort Campbell, Kentucky: Ph.D. Dissertation, College of Earth and Mineral Sciences, The Pennsylvania State University, State College, PA, 254 p. VESPER, D. J.; LOOP, C. M.; AND WHITE, W. W., 2003, Contaminant transport in karst aquifers: Speleogenesis and Evolution of Karst Aquifers, Vol. 1, No. 2, 11 p. VYSOTSKY, Y. B.; KARTASHYNSKA, E. S.; VOLLHARDT, D.; AND FAINERMAN, V. B., 2020, Surface pKa of saturated carboxylic acids at the air/ water Interface: A quantum chemical approach: The Journal of Physical Chemistry, Vol. 124, No. 25, pp. 13809–13818. https://doi .org/10.1021/acs.jpcc.0c03785. WADE, L. G., 2006, Organic Chemistry, 6th ed.: Pearson PrenticeHall, Upper Saddle River, NJ, 1192 p. WANG, Y.; WANG, F.; XIANG, L.; BIAN, Y.; WANG, Z.; SRIVASTAVA, P.; JIANG, X.; AND XING, B., 2022, Attachment of positively and negatively charged submicron polystyrene plastics on nine typical soils: Journal of Hazardous Materials, Vol. 431. https://doi.org/10.1016/ j.jhazmat.2022.128566. WARD, J. W.; WARDEN, J. G.; BANDY, A. M.; FRYAR, A. E.; BRION, G. M.; MACKO, S. A.; ROMANEK, C. S.; AND COYNE, M. S., 2016, Use of nitrogen-15-enriched Escherichia coli as a bacterial tracer in karst aquifers: Groundwater, Vol. 54, No. 6, pp. 830–839. https://doi.org/10.1111/gwat.12426.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
167
Riddell, Vesper, and McDonald WEST VIRGINIA SPELEOLOGICAL SURVEY (WVASS), 2019, The Caves and Karst of Monroe County, West Virginia: WVaSS Barrackville, WV, 408 p. WHITE, W. B.; HERMAN, J. S.; HERMAN, E. K.; AND RUTIGLIANO, M., 2018, Advances in Karst Science: Karst Groundwater Contamination and Public Health Beyond Case Studies: Springer International Publishing, Cham, Switzerland, 329 p. WOESSNER, W. W. AND POETER, E. P., 2020, Hydrogeologic Properties of Earth Materials and Principles of Groundwater Flow: The Groundwater Project, Guelph, Ontario, Canada, 226 p. Electronic document, available at https://gw-project.org/books/
168
hydrogeologic-properties-of-earth-materials-and-principles-ofgroundwater-flow/ XIA, G. AND BALL, W. P., 1999, Adsorption-partitioning uptake of nine low-polarity organic chemicals on a natural sediment: Environmental Science and Technology, Vol. 33, No. 2, pp. 262–269. https://doi.org/ 10.1021/es980581g. ZHOU, Y.; LIU, X.; AND WANT, J., 2019, Characterization of microplastics and the association of heavy metals with microplastics in suburban soil of central China: Science of the Total Environment, Vol. 694, pp. 1–10. https://doi.org/10.1016/j.scitotenv .2019.133798.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 157–168
Potential of Cave Sediments as a Proxy for Tropical Cyclone and Storm Activity JASON POLK* Department of Earth, Environmental, and Atmospheric Sciences, Western Kentucky University, Bowling Green, KY 42101
PHILIP VAN BEYNEN School of Geosciences, University of South Florida, Tampa, FL 33620
Key Terms: Cave Sediments, Paleoclimate, Tropical Cyclones, Karst Environments, Storm Events ABSTRACT Cave sediments show promise as a proxy for the reconstruction of paleo-storm activity. Here, we present a study of allochthonous sediments from two different caves located in west-central Florida that exhibit high variability in sediment layer thickness and are characterized by mostly alternating organic matter/sand couplets. Both sediment records are well constrained chronologically by 210Pb for Vandal Cave in Citrus County and by 14C for Jennings Cave in Marion County, with »50 and 2,700 years of deposition, respectively. Consequently, the Vandal Cave sediments were used to determine whether historic tropical storms produced changes in stratigraphy. The three thickest layers in Vandal Cave correspond with high-precipitation events between 325 and 500 mm. There are similar sedimentary layers found in Jennings Cave, but the upper sediments representing 50 years of deposition are unfortunately highly compacted due to human traffic in the cave. Episodes of intense deposition were noted, specifically from 1,560 to 1,580 years B.P., when 15 cm of sediment was deposited in eight sand layers, indicating a recurrence interval of »2.5 years for major storms. Results from this study suggest cave sediments in certain geographic settings may serve as accurate proxies for storm activity.
INTRODUCTION Paleo-tempestology research is vital for subtropical regions like Florida. Expanding current knowledge of the regularity with which extreme storms make landfall in certain locations along the Florida Peninsula, located in the southeastern United States and bordered by the *Corresponding author email: jason.polk@wku.edu
Gulf of Mexico on the west and the Atlantic Ocean on the east, may help to determine the risk of particular locales for extreme flooding events. Most sedimentary studies investigating the impacts of climate on the state have relied on lacustrine deposits (Grimm et al., 1993, 2006; Watts and Hansen, 1994; Elsner et al., 2000; Huang et al., 2006; and Das et al., 2013). Many of these lakes are flooded sinkholes, which are a characteristic feature of Florida’s karst environment. A recent study by Wang et al. (2021) provided clear evidence of the depositional characteristics of Hurricane Irma in three sinkhole lake sediment deposits in the Florida Keys. Caves are also traps for primarily allochthonous sediments, since allogenic sediments are less common due to the dissolution of the bedrock as opposed to mechanical weathering (White, 1988), but their potential for recording extreme storm activity has received little attention, particularly in Florida. Consequently, we investigated this potential through a detailed examination of sediment profiles from two caves in west-central Florida. Surface soil, primarily comprised of detrital clay, sand, gravel, organic matter, and other regionally specific clastic materials, is transported into caves by overland flow, streams, gravity, or other natural processes (Panno et al., 2004; Polk et al., 2013). Sediments enter the cave via different entry points including entrances, small fissures and cracks in the bedrock, sinkhole piping, enlarged conduits, and fluvial-groundwater interactions (White, 1988; Palmer, 2007). If left undisturbed, a sequence of terrigenous cave sediment lithofacies can record changes in the landscape above the cave in their layering, as well as processes occurring within the cave (Figure 1). A wide range of analytical methods can be employed to determine the age, origin, and environmental history of sediments found in caves. Depending on the type and thickness of the sediment record and method of analysis, the resulting paleo-environmental record can span from hundreds to millions of years (Granger et al., 2001). An important reason for using cave sediments is the protected environment the cave offers,
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
169
Polk and van Beynen
Surface sediments washed into cave Sediment accumula on Limestone
Figure 1. Model of cave sediment deposition process under vegetation conditions. Layers accumulate from the cave floor upward (as indicated by the black arrows pointing upward) during storm events and remain trapped in the cave for tens to thousands of years.
preventing pedogenesis and the in situ alteration of sediments over long periods of time (Ellwood et al., 1997). Researchers have analyzed cave sediments for various purposes, including biological, archaeological, diagenetic, environmental, geomorphological, and hydrological investigations (e.g., Brinkmann and Reeder, 1995; Springer et al., 1997; Panno et al., 2004; and Polk et al., 2013). Only a handful of studies attempted to investigate storm deposits in caves (Van Gundy and White, 2009; Rubin et al., 2017; Switzer et al., 2020; and Pennos et al., 2021). Consequently, this study used cave sediment records from several caves to expand understanding of how they might help researchers interpret the sedimentary signature with respect to storm events, both ancient and modern. More specifically, the research questions investigated were the following: (1) Do cave sediments record past storm events that can be identified via depositional analysis? (2) Can modern cave sediment deposits provide a calibration for the late-Holocene cave sediment deposition as it relates to tropical cyclone and extreme storm activity? STUDY AREA Vandal Cave, part of the Dames Caves complex, is located in the Withlacoochee State Forest (WSF) on the Citrus Tract in Citrus County, approximately 11 km northwest of Brooksville in west-central Florida (Figure 2). The cave’s morphology allows modern, rapid sedimentation to enter the cave from direct overwash from the surface through several large entrances (Figure 2). Anecdotal information from local cavers and WSF personnel suggests that more than 1 m of sediment was deposited in the cave over the last 40 to 60 years. The cave contains minimal passage and is surrounded by an open area where forest trails intersect, 170
with access on a permit-authorized basis; however, the cave is heavily visited and is part of a complex of nearby caves that are frequented by locals. The main passage is approximately 7 m deep, with two small passages formed along joints heading in the northwest (NW) and south (S) directions. Sediments wash in via a north-northwest, south-southeast sloping entrance and through the large 4-m-long by 3-m-wide collapse opening into the main room, where they are deposited (Figure 2). The cave is hydrologically separated from the water table (vadose passage), with no active, perennial stream flowing through it, so sediments are captured as they wash in and remain there. Jennings Cave is located in Marion County in westcentral Florida approximately 15 km west of Ocala (Figure 2) and 50 km north of Vandal Cave. The cave, privately owned by the Southeast Cave Conservancy, is gated to prevent unauthorized access and destruction. Jennings Cave is located in a forested area, with a few dirt roads nearby and little in the way of development or urbanization. The cave has an 8-m-deep, 2-m-wide vertical entrance that opens into several fracturecontrolled vadose passages totaling about 200 m in length. Minor sediment input is visible through several thin fractures and conduits that extend upward several meters into the cave ceiling (Figure 2). Major sedimentation occurs through the vertical entrance shaft as observed during rain events. Sedimentation on the cave floor is widespread throughout the cave and exists up to several meters in depth before bedrock is reached. The cave drains to the water table and serves as a vadose input to the aquifer, with no perennial water flow in the cave’s passages. Geologic Setting The areas surrounding both Vandal Cave and Jennings Cave contain many karst features, including sinkholes,
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
Record of Storm Activity in Cave Sediments
Figure 2. Study area map of Jennings Cave (Marion County, FL) and Vandal Cave (Citrus County, FL) and respective cave maps for each sediment sampling site.
dry valleys, caves, and interfluvial hills (Reeder and Brinkmann, 1998; Florea and Vacher, 2007). The local geology consists of the fossiliferous, highly karstified Ocala Limestone intermittently covered (»8 to 10 m thick) by the Hawthorne Formation’s undifferentiated
clays and sands, which overlie most of the state’s limestone (Lane and Hoenstine, 1991; Florea, 2006). Additional Pliocene–Pleistocene quartz sands, clayey sands, and clays overlie the Hawthorn strata in some areas with varying thicknesses (White, 1970).
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
171
Polk and van Beynen
Climate The regional climate in west-central Florida is humid subtropical, with an average annual temperature of 22°C and average annual precipitation totaling 1,300 mm (SE Regional Climate Center, 2022). The maximum monthly mean of 33.2°C occurs in July, and the minimum monthly mean of 8.8°C occurs in January. The maximum monthly mean precipitation volume of 190 mm occurs in July, and a minimum monthly mean of 50.8 mm occurs in January (SE Regional Climate Center, 2022). The area exhibits an almost monsoonal-like climate, with a wet summer wherein two thirds of the annual rainfall occurs in average years between June and October, with the occasional hurricane, and a drier winter season from November to May (Winsberg, 2003; Alvarez Zarikian et al., 2005). Records of extreme events, including tropical storms, hurricanes, and anomalously high rainfall events, exist for nearly a century, with varying levels of spatial and temporal resolution. The National Oceanic and Atmospheric Administration (NOAA) National Centers for Environmental Information (NOAA, 2022a) and its affiliated sources (Cooperative Observer Network [COOP], Community Collaborative Rain, Hail, and Snow network [CoCoRAHS]) provide, at minimum, daily to annual rainfall totals for Marion and Citrus Counties since the early 1900s, along with tropical cyclone tracks.
The lack of interaction with the water table and active cave streams was particularly important because this removes the potential of disturbance or reorganization of the sediments post-deposition. Sediment Analysis Sediment cores were collected in 2007 for Jennings Cave (JC-07) and in 2008 for Vandal Cave (VC-08). Sediment samples were collected using 10-cm-diameter, schedule 40 polyvinyl chloride (PVC) cores vertically inserted into the sediment sequence in both caves. The core diameter and PVC thickness reduced the possibility of compaction during insertion. Pits were then dug beside the cores to determine whether the layering was proximally consistent around the cores. The cores were then removed laterally into the pit and capped to prevent sediment loss after extraction. The Jennings core length was 110 cm, and the Vandal core was 74 cm. The cores were analyzed at the University of South Florida’s Soils and Physical Geography Laboratory. Cores were cut in half lengthwise, with one half was used for stratigraphic analysis, including description of color, layer properties, grain type, texture, and any other discernible properties (Supplemental Material 1 and 2, https://www.aegweb.org/ e-eg-supplements). The other halves were used for sediment dating.
Soil and Vegetation
Lead-210 (210Pb) Dating
Vegetation around Vandal Cave and Jennings Cave consists mainly of flatwood and mixed hardwood forests, common in this area of Florida (Watts and Collins, 2008). This type of environment includes longleaf pine (Pinus palustris), slash pine (Pinus elliottii), turkey oak (Quercus laevis), live oak (Quercus virginiana), saw palmetto (Serenoa repens), wire grass (Aristida sp.), erycids, species of holly (Ilex), forbs, and various scrub vegetation. Some deforestation has occurred in areas where isolated homes exist, but the area directly around the caves has no human occupation. Soil cover mostly consists of leaf litter and detrital material on the surface, with medium to fine loamy sand beneath that, and the occasional clay layer with silty organics. The common soil type is the Candler fine sand series (Watts and Collins, 2008).
Due to the expectation that the Vandal Cave sediments were modern, 210Pb dating was utilized to date the sediment, with 137Cs used as an independent tracer to validate the age model. Ten evenly spaced individual layers were sampled from the first half of the VC-08 sediment core, weighed to obtain their wet weight, dried overnight in an oven at 40°C, and reweighed to obtain their dry weight. The bulk density was calculated for each sample, and then they were ground using a pestle and mortar, bagged, sealed, and sent to Micro Analytical (subsidiary of BETA Analytic) in Miami, FL, for 210Pb dating.
METHODOLOGY Vandal Cave and Jennings Cave were selected because they met the following criteria: sediment inputs from the surface (entrance, fractures, conduits, etc.), free of bioturbation, vadose passages with no direct interaction with the aquifer or in cave streams (being hydrologically separate from the phreatic zone), and having layered deposition. 172
Radiocarbon Dating For the Jennings Cave core, each 1-cm-thick layer was examined for the presence of sufficient organic carbon in the form of charcoal, seeds, wood, and/or organic matter to use for establishing a chronological record of deposition for the sediment core through radiocarbon dating. Amounts of 0.5 g or more were collected from layers containing sufficient material and then sent to BETA Analytic in Miami, FL, for radiocarbon dating using accelerator mass spectrometry. Dates were calibrated to calendar ages using the CALIB 5.0.1 program and the INTCAL 04 radiocarbon database (Talma and Vogel, 1993; Stuiver et al., 1998; and IntCal04, 2004).
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
Record of Storm Activity in Cave Sediments
Figure 3. VC-08 sediment core and age model. Age-depth plot for VC-08 from 210Pb dates. Error bars define 1 standard deviation; 137Cs is shown versus depth and was used in the CRS age model, with the shift at 25 cm indicated by an arrow. Individual storms of distinction from the historic record are notated on the core alongside depth, providing a reference for the presence of distinct wash-in events from large storms present throughout the core. Reference weather data are from https://weatherspark.com/y/16905/Average-Weather-in-Brooksville-Florida-United-StatesYear-Round.
RESULTS AND DISCUSSION Vandal Cave The sediment core VC-08 from Vandal Cave represents the upper portion of the accumulated sediments in the main passage of the cave. Several test pits were dug throughout the main room to expose the sediment layering, revealing consistent strata across the entire sediment deposit (Figure 3). The PVC core diameter of 10 cm and sample interval of 1 cm provided layer-averaged values for the analyses and eliminated any minute intra-core discrepancies in sediment deposition. Analyses were limited to a physical description and comparison to historic weather data, since this core was meant to be a modern, high-resolution proxy in an attempt to calibrate the JC-07 core data with regard to climate and sedimentation processes. Vandal Cave Sediment Physical Characteristics The physical description of the VC-08 core (see Supplemental Material 1, https://www.aegweb.org/e-egsupplements) provides generalized information on each 1-cm-thick layer, including sublayers, about the color,
physical characteristics (grain type, size, composition), and groupings of similar layers (i.e., multiple-centimeter sections that comprise a layer unit). The physical description illustrates the variability of the sediment layering and the episodic nature of its deposition. Primarily, the core consisted of recurring sand and organic layers, somewhat rhythmic (e.g., varved), although not consistently alternating throughout the entire depth of the core (Figure 3; see Supplemental Material 1, https://www.aegweb.org/e-eg-supplements). There was evidence of previous anthropogenic presence in the cave, with remnants of glass, plastic, and metal present in a few of the upper layers in minimal quantities. The VC-08 sediment core consisted of 19 organic matter layers (comprised of silty, clayey, organic-rich sand), 11 mixed sandy organic layers, 3 orange ironstained clayey-sand layers, 18 pure, fine sand layers, 1 very fine sand layer, 1 gray, charcoal layer, and 1 black, unconsolidated, sandy, old floor layer, all of varying thicknesses and horizon types, ranging from abrupt to wavy (Figure 3). Overall, the entire core consisted mostly of quartz sand and interspersed clayey sands, based on analysis under a microscope and absence of fizzing under HCl testing, indicating a lack of carbonate
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
173
Polk and van Beynen
materials. Wood (1996) performed textural and grainsize analyses for a sediment bank from Vandal Cave, but little of the data were useful for delineating storm event or climatic variability solely from their physical sediment data and low sampling resolution. Vandal Cave 210Pb Dating Results Bulk sediment lead-210 (210Pb) dates from 10 layers provided a chronology of sediment deposition for the VC-08 sediment core (Table 1). This geochronological dating method was used because of the expected young age of the sediments, and it provides greater dating precision for materials less than 150 years old. Table 1 shows the dates with errors corresponding to core depths where they were obtained. A simple linear age model was constructed based on the constant rate of supply (CRS) age calculation model (Aquino-López et al., 2020) used for the 210Pb dating (Figure 3). This model implies a constant sediment flux with constant initial 210 Pb concentrations but does not require a constant rate of sedimentation. The presence of 137Cs in the sediments indicates that all sediments are younger than 1963 6 2 years A.D., when nuclear weapons testing began (Schelske and Hodell, 1995). The dating model supports a relatively constant sedimentation rate until about 25 cm, when a spike in the 137Cs and associated shift in deposition rate occurred, with the dating model supporting a much faster rate of sedimentation occurring after the calendar year 2000. Vandal Cave Sediments Discussion The analysis of the sediment core from Vandal Cave (VC-08) provides insight on the mechanism of recent, shortterm sediment deposition from approximately 1964 to 2008 in a Florida cave. It also illustrates the complexity of the ever-changing surface environment and corresponding processes influencing sediment deposition in caves. Rapid sedimentation rates and accumulations for Vandal Cave can be attributed to the large unroofed entrance, which allows direct, unobstructed sedimentation to occur in the terminal passage of the cave. Sediment transport is likely only over a short distance, deriving from allochthonous sources adjacent to the cave entrance, the abundance of open, sandy areas nearby, the angular sand grains, and personal observation of sediment movement during storms, thereby enhancing the opportunity of sediment entrainment and mobilization before deposition (Figure 2). Additionally, the cave is the lowest topographic point on the landscape and drains a large area, thereby increasing the potential for sediment input during severe events when storm intensity would generate runoff velocities that could carry in the largest grain size (sand), and likely highest volume, of sediment for rapid deposition (Figure 4). 174
One easily discernible characteristic of the VC-08 sediment core is its prominent layering (Figure 3). Based on dating of the sediment layers, almost half (»30 cm) of the sediment was deposited between 2000 and 2008 (Supplemental Material 1, https://www.aegweb.org/e-egsupplements). The remaining 44 cm section of sediment was deposited in the 36 years from 1964 to 2000, indicating variable rates of deposition that could be attributed to changes in precipitation. The physical description indicates multiple thick sand layers, which can be attributed to highprecipitation events (Figures 5 and 6). Some of these layers correspond with known major storms, such as Hurricane Elena, but they may also have been deposited as a result of local, extreme thunderstorms that produced tens of centimeters of precipitation within several hours. Despite the different origins of severe storms, the same erosive mechanism of significant overland flow applies for both localized thunderstorms and cyclones. During the 44 years of sedimentation, only five such events were documented. Other, thinner sand layers are also present in the core, and the rest of the deposits suggest that more regular, smaller precipitation events were the more significant contributors to cave sedimentation. An analysis of annual precipitation during depositional periods in core VC-08 indicated several periods of increased precipitation corresponding to increased sedimentation and thick sand deposits (Figure 5), along with those distinct events described in Figure 3, which suggest a relationship between intense storms and sediment layers. For example, several large storm events (February 2005, December 2002, Hurricane Elena in 1985, etc.) occurred during periods of thicker sand deposits in the VC-08 core. These periods also coincide with wetter-than-average periods (multiyear) seen in the precipitation record for the area (Figure 5), thereby suggesting the cave sediment record serves as a robust proxy for periods of more intense storms that cause higher-than-average annual rainfall amounts. Couplets caused by wet season deposition, as suggested by Brinkmann and Reeder (1995) and Wood (1996), are not continuous for every year in the VC-08 sediment core, as too many layers exist to account for seasonal couplets for every year of sediment deposition. Thus, a more plausible explanation is that an event must be sufficiently intense to generate surface runoff; therefore, hiatuses or interannual variations exist within the core, with couplet deposition driven by storm events and seasonality. Larger, more frequent summer storms and strong winter low-pressure systems could both cause increased sand erosion and transport, resulting in layered, coupled deposition in the cave irrespective of season. Often, wetter years in Florida are due to a combined higher winter precipitation and average summer amounts, as with El Niño phases, which would cause integrated mixing of sandy and organic layers to occur during severe winter frontal events and summer storms (Donders et al., 2005).
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
1.5 2.5 15.5 36.5 51.5 63 82.5 100.5 104
BETA-228643 BETA-228644 BETA-228645 BETA-228656 BETA-28647 BETA-228648 BETA-228649 BETA-228650 BETA-228651
Charcoal Charcoal Wood Charcoal Organics Charcoal Charcoal Organics Jaw Bone
Description Post-bomb 370 610 1,560 1,610 1,930 2,210 2,680 2,650
Age (14C year B.P.)
42.36 63.43 83.17 54.01 49.09 46.58 56.47 65.39 70.33 64.8
Wet Weight (g) 29.6 49.19 57.75 43.98 43.38 40.13 42.05 59.34 65.88 46.59
Dry Weight (g) 1.79 1.99 1.78 2.08 2.26 2.21 1.91 2.32 2.4 1.84
Rho
Pb-214 (dpm/g) 2.1 6 0.27 2.37 6 0.16 2.17 6 0.12 1.59 6 0.14 1.69 6 0.13 1.18 6 0.14 1.95 6 0.15 0.98 6 0.12 1.08 6 0.12 1.89 6 0.16
Pb-210(total) (dpm/g) 3.2 6 0.38 2.82 6 0.21 2.69 6 0.2 2.27 6 0.19 1.86 6 0.18 1.93 6 0.19 3.64 6 0.23 1.52 6 0.16 1.31 6 0.14 3.7 6 0.23
Pb dates are given in calendar years. CRS ¼ constant rate of supply.
1 5 10 15 25 30 40 50 65 74
MA-2819 MA-2820 MA-2821 MA-2822 MA-2823 MA-2824 MA-2825 MA-2826 MA-2827 MA-2828
210
Depth (cm)
Sample No.
(B) VC-08 Bi-214 (dpm/g) 2.48 6 0.26 2.39 6 0.14 1.91 6 0.12 1.7 6 0.12 1.71 6 0.11 1.38 6 0.13 1.98 6 0.13 0.62 6 0.1 1.03 6 0.08 1.73 6 0.14
0.4 pMC 40 40 40 40 40 40 40 40
6 (1r)
Accelerator mass spectrometry radiocarbon dates from JC-07 are given in calendar year B.P. (2008).
Depth (cm)
Sample ID
(A) JC-07
0.91 6 0.3 0.44 6 0.17 0.656 0.15 0.62 6 0.15 0.15 6 0.14 0.65 6 0.15 1.68 6 0.17 0.62 6 0.12 0.25 6 0.11 1.89 6 0.17
Pb-210(excess) (dpm/g)
50 340 600 1,560 1,580 1,820 2,210 2,670 2,710
Age (calendar year B.P.)
Table 1. Age model data for the (A) JC-07 (14C) and (B) VC-08 (210Pb) sediment cores.
2008 6 1 2007 6 1 2005 6 1 2003 6 1 2002 6 1 2000 6 2 1991 6 3 1979 6 7 1971 6 11 1964 6 14
CRS(year)
— 40 40 40 40 40 40 40 40
6 (2r)
.0.05 6 — 0.24 6 0.06 0.28 6 0.06 0.23 6 0.06 2.98 6 0.11 0.23 6 0.06 0.39 6 0.07 0.22 6 0.05 0.2 6 0.06 1.02 6 0.09
Cs-137 (dpm/g)
1.31 6 0.03 1.96 6 0.05 2.32 6 0.05 0.46 6 0.01 0.4 6 0.01 0.5 6 0.01 1.9 6 0.04 0.4 6 0.01 0.85 6 0.02 2.65 6 0.06
K-40 (dpm/g)
1958 CE 1668 CE 1408 CE 448 CE 428 CE 188 CE 202 BCE 662 BCE 702 BCE
Age (calendar year CE before 2008)
Record of Storm Activity in Cave Sediments
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
175
Polk and van Beynen
Figure 4. GTOPO60 (60 arc-second digital elevation model) map of Vandal Cave. Map shows its topography, which is ideal to drain the landscape and rapidly accumulate sediment (adapted from Citrus County geographic information system).
Jennings Cave The sediment core JC-07 from Jennings Cave was representative of the entire accumulation of sediment in this section of the main passage of the cave. This conclusion is based on a trench dug laterally across the passage, which showed consistency in sediment deposition and stratigraphy to bedrock. From the physical description, the variability of the sediment layering and the episodic nature of its deposition were evident (Supplemental Material 2, https:// www.aegweb.org/e-eg-supplements). Primarily, the core consisted of recurring sand and organic layers, somewhat rhythmic, or varved, in nature, although not consistently alternating throughout the entire depth of the core (Figure 7; see Supplemental Material 2, https://www.aegweb.org/e-egsupplements). Grain-size sieve analysis showed the sediment core consisted of 64 fine sand layers, 69 organic matter layers (comprised of silty, clayey, organic-rich sand), 16 mixed sandy organic layers, and 2 orange iron-stained clayey-sand layers, all of varying thicknesses and fairly abrupt horizons (Polk et al., 2013). Overall, the entire core consisted mostly of quartz sand, based on analysis under a microscope and absence of fizzing under HCl testing, indicating a lack of carbonate materials. Sand was present within every layer in varying amounts and is expected in west-central Florida, where Quaternary quartz sand dominates the soil matrix (Watts and Collins, 2008). Prior textural 176
and grain-size sieve analyses indicated minimal physical variation within the layers of the core (Polk et al., 2013). Jennings Cave Radiocarbon Dating Radiocarbon (14C) dates from charcoal, wood, seeds, and organic matter, and a rabbit jawbone were determined from nine layers of sediment from Jennings Cave core JC-07 (Table 1; Figure 7). The nine age determinations provided a chronology of sediment deposition that exhibits fairly consistent sedimentation rate over a period of »2,700 years. A depth-to-age model was constructed using a fourth-order polynomial regression (r2 ¼ 0.987, p , 0.001) model to provide a chronological record for the sediment deposition (Figure 7; Polk et al., 2013). This regression model was chosen to align the best fit within the depth-age model while using the fewest terms possible within the given age errors. The number of dates used (nine) and their low errors provided acceptable confidence in the timescale reconstruction for this study as compared to similar methodologies used in lacustrine and marine sediment studies (Curtis et al., 1998, 1999; Hodell et al., 2005). Jennings Cave Discussion One important consideration of the JC-07 sediments is how they accumulated in the cave to form the layers
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
Record of Storm Activity in Cave Sediments
Figure 5. Total annual rainfall for Citrus County by year (1950–2008). Dark blue lines indicate higher-than-average rainfall (1,376 mm/yr for study period, indicated by light blue cutoff), and red brackets indicate prolonged periods of rainfall above average (NOAA, 2022a).
seen in Figure 7. The geomorphology of the cave is maze-like and extensively fracture-controlled, with a single 10-m-deep vertical entrance (Figure 2). From personal observation, sedimentation occurs mainly through the entrance during wash-in events and is present throughout the entire cave at a fairly uniform, horizontal level. Nearly all of the core comprised couplets of fine organic matter and coarse sand. It is likely that major storm events would have been necessary to meet the conditions of the Hjulström curve (Selley, 1982) required to transport the sand, silt, clay, and organic matter into the cave to deposit the layers seen in the core (Figure 6). Using the interpretation from Vandal Cave, we propose that surface soils were washed into the cave primarily during major precipitation events, as there are only 69 couplets in the core over an »2,700 period. Table 2 provides a more detailed analysis of the JC-07 layering regarding the frequency of such precipitation events using the chronology and occurrences of sand deposits (Supplemental Material 2, https://www.aegweb.org/ e-eg-supplements). Sedimentation rates are fairly consistent throughout the JC-07 record, except for a period of extreme deposition between 1560 and 1580, where eight sand layers are present. There is no evidence of abundant charcoal in this portion of the core, which rules out the removal of vegetation above the cave, allowing for enhanced soil erosion, nor would local thunderstorms be the likely cause, because a more regular occurrence of
these events would be expected over 2,700 years; hence, there would be more couplets. While there is little evidence from the regional paleo-tempestology records of such a period of intense hurricane activity, there are few such records south of Florida’s panhandle (Ercolani et al., 2015; Yao et al., 2020). During the last »350 years, there were few storms, suggesting a period of cooler sea-surface temperatures in the tropical North Atlantic Ocean (Lane and Donnelly, 2012) inhibiting atmospheric-oceanic circulations (van Hengstum et al., 2016) or diversion of the storms elsewhere (McCloskey et al., 2013). Cave Sediments as Records of Climate and Other Extreme Events Many studies have investigated allochthonous cave sediments as records of speleogenesis (Granger et al., 2001; Farrant and Smart, 2011; Martini, 2011; Springer et al., 2014; and Laureano et al., 2016) and long-term climate change (Panno et al., 2004; Polk et al., 2007, 2013; Aidona et al., 2013; Benedetti et al., 2019; and Hajna et al., 2021); however, only a few have used clastic sediments to investigate abrupt, extreme events. Pennos et al. (2021) observed that sediments deposited from flooding in a cave in Greece impacted human occupation of the site. The sediments were deposited from a river flowing close to the cave entrance. Several other studies documented changes in passage sedimentation from cave stream fluctuations or
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
177
Polk and van Beynen
Figure 6. Hurricane and tropical storm tracks across the study area from 1950 to 2008 from NOAA HURDAT2 geographic information system data (NOAA, 2022b) overlapping the period of deposition for the Vandal Cave sediment core. Several storms correspond to noticeable layers of thick sand deposits in the core (e.g., 1971, 1974, 1985, 2002–2004). General locations of Jennings Cave (JC) and Vandal Cave (VC) are noted by red and blue dots, respectively.
backwater flooding from surface rivers adjacent to cave entrances (Doehring and Vierbuchen, 1971; Springer and Kite, 1997; and Herman et al., 2008), but the study by Van Gundy and White (2009) is the only one in the United States to measure the impact of storms on cave sediments resulting in depositional sequences without existing cave stream influences, though in their particular case, the storm flushed sediments from the cave. Evidence remained of the sediments, with deposits being trapped within certain portions of the cave passages; yet, none of these studies had deposition originating from sediments entrained by surface runoff adjacent to the cave with no input from pre-existing water bodies (streams/rivers). It should be noted that not all clastic sediments that record extreme events are a result of extreme precipitation events, as is the situation with our Florida study caves. Switzer et al. (2020) determined that well-preserved 178
overwash sediments in a coastal cave closely resembled those of the nearshore environment. Their interpretation was that these sediments were archives of paleo-storm activity. Coastal caves have also been shown to act as depositories from tsunamis that occurred over the last 7,400 years in the Indian Ocean (Rubin et al., 2017). In our study, the upland setting provides an ideal environment for sediments to be introduced only through runoff deposition, with protection from disturbance due to the vadose nature and high elevation of the caves. Severe storms of high intensity and low duration and frequency typically produce runoff and, in turn, can cause higher-than-average hydrologic year rainfall accumulation. Increased runoff and sedimentation are products of wetter seasons, which may result in the higher deposition of sandy sediment and higher number of sediment couplets during those periods. Sustained climate anomalies
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
Record of Storm Activity in Cave Sediments
Figure 7. Jennings Cave JC-07 sediment core and age model. Age-depth plot for JC-07 from radiocarbon dates over 110 cm of deposition.
causing warmer, wetter conditions with an increased frequency of tropical cyclones also will increase sedimentation and should be recorded in cave deposits under the described conditions (being protected from external inputs and undisturbed). For example, average sediment thickness in VC-08 was 7.3 mm/layer, whereas average thickness in JC-07 was 4.4 mm/layer, suggesting a much higher amount of sediment deposition per event in Vandal Cave, likely due to its open entrance, allowing more direct input of sediment under less severe storms than
Jennings Cave. Jennings Cave, in turn, would likely record longer-scale climate impacts of wetter periods due to its morphology, as shown by thicker sediment accumulation and more couplets (Table 2), which is ideal for garnering paleo-climate interpretations from cave sediments. ASSUMPTIONS AND LIMITATIONS OF STUDY Assumptions of this study include regular sediment deposition driven by runoff into the caves, calibration of
Table 2. Sediment deposition and characteristics data for the JC-07 core from Jennings Cave. Age of Sediments (cal year before 2008) 0–50 50–340 340–600 600–1,560 1,560–1,580 1,580–1,820 1,820–2,210 2,210–2,670 2,670–2,710
Time Interval (years)
Sediment Thickness (mm)
Sedimentation Rate (mm/yr)
Number of Sand Layers
Recurrence Period of Storms (years)
50 290 260 960 20 240 390 460 40
15 10 13 21 15 11.5 19.5 18 3.5
0.3 0.03 0.05 0.02 0.75 0.05 0.05 0.04 0.09
0 2 11 13 8 9 12 11 0
* 145 24 46 2.5 27 32 42 —
*The high visitation of the cave by humans over the last 50 years has led to much compaction and disturbance of the upper sediments. Consequently, this section in not included in the discussion.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
179
Polk and van Beynen
dates based on given errors, lack of alteration of core sediments prior to sampling, and sedimentation size distribution that follows Hjulström’s curve driven by runoff velocity. The differences between rates of deposition, anthropogenic disturbance, and sediment deposition rate data reveal a much more complex system and response to external factors during the short-term period. Also, the size of the cave passage and distribution of the washed-in sediment could affect the thickness of layers, with the overall average in Jennings Cave being lower than that in Vandal Cave due to larger floor surface area. These factors contribute considerably to the interpretation of the different sediment data for both caves, requiring a more individualized approach to examining and understanding the influences affecting each cave’s sediment record. CONCLUSIONS Layering in cave sediments is a product of multiple complex influences that include episodic deposition from storm events, settling mechanisms, post-depositional anthropogenic processes, and longer-term seasonal and climatic influences. To revisit the possibility of using the data from VC-08 as a modern calibration and interpretation for the environmental and climatic processes affecting the JC-07 sediment core, several factors must be considered. Sediment deposition in Vandal and Jennings Caves, while not an entirely stochastic process, is episodic in nature and demonstrates the usefulness of using sediment deposition for interpreting stormand climate-driven events in a hydrologically closed cave with a sensitive and direct connection to the surface. Changes in vegetation can be dynamic and rapid, on the order of 100 years or less, and more intense due to extreme climatic events (Boutton et al., 1998); however, the temporal span of the VC-08 sediment core, approximately 44 years, is likely too short to encompass entire shifts in vegetation regimes as dramatic as that which can occur during major climatic events, such as the Little Ice Age of the 16th to 19th centuries or more protracted glacial-interglacial periods (Krull et al., 2005). In contrast, the JC-07 core spans a climatically significant period over 2,700 years and records several abrupt, prolonged climate anomalies, with some couplet layers likely being indicative of intense rainfall events or years. However, identifying individual events is difficult due to the rate and amount of sedimentation, but using the Vandal Cave sediment core as a calibration proxy proved to be useful to interpret the variation within the longer JC-07 record with respect to thicker sand layering and more frequent couplets indicative of wetter periods. 180
ACKNOWLEDGMENTS We would like to thank the Florida Cave Conservancy and the Southeastern Cave Conservancy for allowing access to Jennings Cave for coring and sample collection. Thanks go to Colleen Werner and Withlacoochee State Forest for access and permission to collect core samples at Vandal Cave. To those who assisted with the coring and fieldwork, including Grant Harley, Tom Turner, Robert Brooks, and Jon Sumrall, as well as those who helped with sediment sample processing, including Travis Stull, Chris Lizzardi, Emily Wakely, and Ray Vinson, we convey our thanks. Much appreciation goes to Drs. Philip Reeder and Bob Brinkmann for their discussion and insights on the past sedimentation at Vandal Cave. Special thanks go to Brittany Pekara and Meghan Forbes for assistance with geographic information systems and figure development. This study was funded in part by the Geological Society of America and Southwest Florida Water Management District. REFERENCES AIDONA, E.; PECHLIVANIDOU, S.; AND PENNOS, C., 2013, Environmental magnetism: Application to cave sediments: Bulletin of the Geological Society of Greece, Vol. 47, No. 2, pp. 892–900. ALVAREZ ZARIKIAN, C. A.; SWART, P. K.; GIFFORD, J. A.; AND BLACKWELDER, P. L., 2005, Holocene paleohydrology of Little Slat Spring, Florida, based on ostracod assemblages and stable isotopes: Palaeogeography, Palaeoclimatology, Palaeoecology, Vol. 225, No. 1–4, pp. 134–156. AQUINO-LÓPEZ, M. A.; SANDERSON, N. K.; BLAAUW, M.; SANCHEZCABEZA, J. A.; RUIZ-FERNANDEZ, A. C.; AQUINO-LÓPEZ, J.; AND CHRISTEN, J. A., 2020, A simulation study to compare 210Pb dating data analyses: arXiv, 2012.06819. BENEDETTI, M. M.; HAWS, J. A.; BICHO, N. F.; FRIEDL, L.; AND ELLWOOD, B. B., 2019, Late Pleistocene site formation and paleoclimate at Lapa do Picareiro, Portugal: Geoarchaeology, Vol. 34, No. 6, pp. 698–726. BOUTTON, T. W.; ARCHER, S. R.; MIDWOOD, A. J.; ZITZER, S. F.; AND BOL, R., 1998, d13C values of soil organic carbon and their use in documenting vegetation change in a subtropical savanna ecosystem: Geoderma, Vol. 82, No. 1–3, pp. 5–41. BRINKMANN, R. AND REEDER, P., 1995, The relationship between surface soils and cave sediments: An example from west central Florida, USA: Cave and Karst Science, Vol. 22, pp. 95–102. CURTIS, J. H.; BRENNER, M.; AND HODELL, D. A., 1999, Climate change in the Lake Valencia Basin, Venezuela, »12,600 yr BP to present: The Holocene, Vol. 9, No. 5, pp. 609–619. CURTIS, J. H.; BRENNER, M.; HODELL, D. A.; BALSER, R. A.; ISLEBE, G. A.; AND HOOGHEIMSTRA, H., 1998, A multi-proxy study of Holocene environmental change in the Maya Lowlands of Peten, Guatemala: Journal of Paleolimnology, Vol. 19, pp. 139–159. DAS, O.; WANG, Y.; DONOGHUE, J.; XU, X.; COOR, J.; ELSNER, J.; AND XU, Y., 2013, Reconstruction of paleostorms and paleoenvironment using geochemical proxies archived in the sediments of two coastal lakes in northwest Florida: Quaternary Science Reviews, Vol. 68, pp. 142–153. DOEHRING, D. O. AND VIERBUCHEN, R. C., 1971, Cave development during a catastrophic storm in the Great Valley of Virginia: Science, Vol. 174, No. 4016, pp. 1327–1329.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
Record of Storm Activity in Cave Sediments DONDERS, T. H.; WAGNER, F.; AND VISSCHER, H., 2005, Quantification strategies for human-induced and natural hydrological changes in wetland vegetation, southern Florida, USA: Quaternary Research, Vol. 64, pp. 333–342. ELLWOOD, B. B.; PETRUSO, K. M.; AND HARROLD, F. B., 1997, Highresolution paleoclimatic trends for the Holocene identified using magnetic susceptibility data from archaeological excavations in caves: Journal of Archaeological Science, Vol. 24, pp. 569–573. ELSNER, J. B.; LIU, K. B.; AND KOCHER, B., 2000, Spatial variations in major US hurricane activity: Statistics and a physical mechanism: Journal of Climate, Vol. 13, No. 13, pp. 2293–2305. ERCOLANI, C.; MULLER, J.; COLLINS, J.; SAVARSE, M.; AND SQUICCIMARA, L., 2015, Intense southwest Florida hurricane landfalls over the past 1000 years: Quaternary Science Reviews, Vol. 126, pp. 17–25. FARRANT, A. R. AND SMART, P. L., 2011, Role of sediment in speleogenesis; sedimentation and paragenesis: Geomorphology, Vol. 134, No. 1–2, pp. 79–93. FLOREA, L. J., 2006, Architecture of air-filled caves within the karst of the Brooksville Ridge, west-central Florida: Journal of Cave and Karst Studies, Vol. 68, No. 2, pp. 64–75. FLOREA, L. J. AND VACHER, H. L., 2007, Eogenetic karst hydrology: Insights from the 2004 hurricanes, peninsular Florida: Groundwater, Vol. 45, No. 4, pp. 439–446. GRANGER, D. E.; FABEL, D.; AND PALMER, A. N., 2001, Pliocene-Pleistocene incision of the Green River, Kentucky, determined from radioactive decay of cosmogenic 26Al and 10Be in Mammoth Cave sediments: Geological Society of America Bulletin, Vol. 113, No. 7, pp. 825–836. GRIMM, E. C.; JACOBSON, G. L.; WATTS, W. A.; HANSEN, B. C. S.; AND MAASCH, K. A., 1993, A 50,000 year record of climate oscillations from Florida and its temporal correlation with the Heinrich events: Science, Vol. 261, No. 5118, pp. 198–200. GRIMM, E. C.; WATTS, W. A.; JACOBSEN G. L., Jr.; HANSEN, B. C. S.; ALMQUIST, H. R.; DIEFFENBACHER-KRALL, A. C., 2006, Evidence for warm wet Heinrich events in Florida: Quaternary Science Reviews, Vol. 25, pp. 2197–2211. HAJNA, N. Z.; MIHEVC, A.; BOSAK, P.; PRUNER, P.; HERCMAN, H.; HORÁČEK, I.; WAGNER, J.; ČERMÁK, S.; PAWLAK, J.; SIERPIEŃ, P.; AND KDÝR, Š., 2021, Pliocene to Holocene chronostratigraphy and palaeoenvironmental records from cave sediments: Ra ciška pe cina section (SW Slovenia): Quaternary International, Vol. 605, pp. 5–24. HERMAN, E. K.; TORAN, L.; AND WHITE, W. B., 2008, Threshold events in spring discharge: Evidence from sediment and continuous water level measurement: Journal of Hydrology, Vol. 351, No. 1–2, pp. 98–106. HODELL, D. A.; BRENNER, M.; AND CURTIS, J. H., 2005, Terminal classic drought in the northern Maya lowlands inferred from multiple sediment cores in Lake Chichancanab (Mexico): Quaternary Science Reviews, Vol. 24, pp. 1413–1427. HUANG, Y.; SHUMAN, B.; WANG, Y.; WEBB, T.; GRIMM, E. C.; AND JACOBSON, G. L., 2006, Climatic and environmental controls on the variation of C3 and C4 plant abundance in central Florida for the past 62,000 years: Palaeogeography, Palaeoclimatology, Palaeoecology, Vol. 237, pp. 428–435. INTCAL04, 2004, Intcal04 terrestrial radiocarbon age calibration, 0– 26 cal kyr BP: Radiocarbon, Vol. 46, No. 3, pp. 1029–1058. KRULL, E. S.; SKJEMSTAD, J. O.; BURROWS, W. H.; BRAY, S. G.; WYNN, J. G.; BOL, R.; SPOUNCER, L.; AND HARMS, B., 2005, Recent vegetation changes in central Queensland, Australia: Evidence from d13C and 14C analyses of soil organic matter: Geoderma, Vol. 126, No. 3–4, pp. 241–259. LANE, E. AND HOENSTINE, R. W., 1991, Environmental Geology and Hydrogeology of the Ocala Area, Florida: Special Publication 31, Florida Geological Survey, Tallahassee, FL, 71 p.
LANE, P. AND DONNELLY, J. P., 2012, Hurricanes and typhoons: Will tropical cyclones become stronger and more frequent: PAGES Newsletter, Vol. 20, No. 1, pp. 32–33. LAUREANO, F. V.; KARMANN, I.; GRANGER, D. E.; AULER, A. S.; ALMEIDA, R. P.; CRUZ, F. W.; STRÍCKS, N. M.; AND NOVELLO, V. F., 2016, Two million years of river and cave aggradation in NE Brazil: Implications for speleogenesis and landscape evolution: Geomorphology, Vol. 273, pp. 63–77. MARTINI, I., 2011, Cave clastic sediments and implications for speleogenesis: New insights from the Mugnano Cave (Montagnola Senese, Northern Apennines, Italy): Geomorphology, Vol. 134, No. 3–4, pp. 452–460. MCCLOSKEY, T. A.; BIANCHETTE, T. A.; AND LIU, K. B., 2013, Track patterns of landfalling and coastal tropical cyclones in the Atlantic basin, their relationship with the North Atlantic Oscillation (NAO), and the potential effect of global warming: American Journal of Climate Change, Vol. 2, pp. 12–22. NOAA (National Oceanic and Atmospheric Administration), 2022a, National Centers for Environmental Information, Climate at a Glance: County Time Series: Electronic document, available at https://www.ncei.noaa.gov/access/monitoring/climate-at-a-glance/ county/time-series NOAA (National Oceanic and Atmospheric Administration), 2022b, National Hurricane Center, HURDAT2 GIS Database: Electronic document, available at https://coast.noaa.gov/hurricanes/#map PALMER, A. N., 2007, Cave Geology: Cave Books, Dayton, OH, 454 p. PANNO, S. V.; CURRY, B. B.; WANG, H.; HACKLEY, K. C.; LIU, C. L.; LUNDSTROM, C.; AND ZHOU, J., 2004, Climate change in southern Illinois, USA, based on the age and d13C of organic matter in cave sediments: Quaternary Research, Vol. 61, pp. 301–313. PENNOS, C.; PECHLIVANIDOU, S.; AIDONA, E.; BOURLIVA, A.; LAURITZEN, S. E.; SCHOLGER, R.; AND KANTIRANIS, N., 2021, Decoding shortterm climatic variations from cave sediments over the Mid-Holocene: Implications for human occupation in the Katarraktes Cave System, northern Greece: Zeitschrift f €ur Geomorphologie, Vol. 63, No. 1, pp. 7–80. POLK, J. S.; VAN BEYNEN, P.; ASMEROM, Y.; AND POLYAK, V., 2013, Reconstructing past climates using carbon isotopes from fulvic acids in cave sediments: Chemical Geology, Vol. 360–361, pp. 1–9. POLK, J. S.; VAN BEYNEN, P.; AND REEDER, P., 2007, Late Holocene environmental reconstruction using cave sediments from Belize: Quaternary Research, Vol. 70, pp. 53–63. REEDER, P. AND BRINKMANN, R., 1998, Paleoenvironmental reconstruction of an Oligocene-aged island remnant in Florida, USA: Cave and Karst Science, Vol. 25, pp. 7–13. RUBIN, C. M.; HORTON, B. P.; SIEH, K.; PILARCZYK, J. E.; DALY, P.; ISMAIL, N.; AND PARNELL, A. C., 2017, Highly variable recurrence of tsunamis in the 7,400 years before the 2004 Indian Ocean tsunami: Nature Communications, Vol. 8. No. 1, pp. 1–12. SCHELSKE, C. L. AND HODELL, D. A., 1995, Using carbon isotopes of bulk sedimentary organic matter to reconstruct the history of nutrient loading and eutrophication in Lake Erie: Limnology and Oceanography, Vol. 40, No. 5, pp. 918–929. SELLEY, R. C., 1982, An Introduction to Sedimentology: Academic Press, New York, 417 p. SE Regional Climate Center, 2022, Southeast Regional Climate Center, retrieved on May 20, 2022 from https://sercc.com/ SPRINGER, G. S. AND KITE, J. S., 1997, River-derived slackwater sediments in caves along Cheat River, West Virginia: Geomorphology, Vol. 18, No. 2, pp. 91–100. SPRINGER, G. S.; KITE, J. S.; AND SCHMIDT, V. A., 1997, Cave sedimentation, genesis, and erosional history in the Cheat River Canyon, West Virginia: Geological Society of America Bulletin, Vol. 109, No. 5, pp. 524–532.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
181
Polk and van Beynen SPRINGER, G. S.; POSTON, H. A.; HARDT, B.; AND ROWE, H. D., 2014, Groundwater lowering and stream incision rates in the Central Appalachian Mountains of West Virginia, USA: International Journal of Speleology, Vol. 44, No. 1, pp. 10. STUIVER, M.; REIMER, P. J.; BARD, E.; BECK, J. W.; BURR, G. S.; HUGHEN, K. A.; KROMER, B.; MCCORMAC, G.; VAN DER PLICHT, J.; AND SPURK, M., 1998, IntCal98 radiocarbon age calibration, 24,000–0 cal BP: Radiocarbon, Vol. 40, No. 3, pp. 1041–1083. SWITZER, A. D.; FELIX, R. P.; SORIA, J. L. A.; AND SHAW, T. A., 2020, A comparative study of the 2013 typhoon Haiyan overwash sediments from a coastal cave and beach system at Salcedo, Eastern Samar, central Philippines: Marine Geology, Vol. pp. 419, 106083. TALMA, A. S. AND VOGEL, J. C., 1993, A simplified approach to calibrating C14 dates: Radiocarbon, Vol. 35, No. 2, pp. 317–322. VAN GUNDY, J. J. AND WHITE, W. B., 2009, Sediment flushing in Mystic Cave, West Virginia, USA, in response to the 1985 Potomac Valley flood: International Journal of Speleology, Vol. 38, No. 2, pp. 103–109. VAN HENGSTUM, P. J.; DONNELLY, J. P.; FALL, P. L.; TOOMEY, M. R.; ALBURY, N. A.; AND KAKUK, B., 2016, The Intertropical Convergence Zone modulates intense hurricane strikes on the western North Atlantic margin: Scientific Reports, Vol. 6, No. 1, pp. 1–10.
182
WANG, Y.; VAN BEYNEN, P.; WANG, P.; BROOKS, G.; HERBERT, G.; AND TYKOT, R., 2021, Investigation of sedimentary records of Hurricane Irma in sinkholes, Big Pine Key, Florida: Progress in Physical Geography: Earth and Environment, Vol. 45, No. 6, pp. 885–906. WATTS, F. C. AND COLLINS, M. E., 2008, Soils of Florida: Soil Science Society of America Inc., Madison, WI, 88 p. WATTS, W. A. AND HANSEN, B. C. S., 1994, Pre-Holocene and Holocene pollen records of vegetation history from the Florida peninsula and their climatic implications: Palaeogeography, Palaeoclimatology, Palaeoecology, Vol. 109, pp. 163–176. WHITE, W. A., 1970, The Geomorphology of the Florida Peninsula: Bulletin No. 51, Florida Bureau of Geology, Tallahassee, FL, 164 p. WHITE, W. B., 1988, Geomorphology and Hydrology of Karst Terrains: Oxford University Press, New York, 464 p. WINSBERG, M. D., 2003, Florida Weather: University of Florida Press, Gainesville, FL. WOOD, H. R., 1996, Recent Sedimentation in Vandal Cave, Citrus County, Florida: Unpublished Honor’s Thesis, University of South Florida, Tampa, FL, 41 p. YAO, Q.; LIU, K. B.; RODRIGUES, E.; BIANCHETTE, T.; ARAGÓN-MORENO, A. A.; AND ZHANG, Z., 2020, A geochemical record of Late-Holocene hurricane events from the Florida Everglades: Water Resources Research, Vol. 56, No. 8, e2019WR026857.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 169–182
Recommended Planning and Response for Hazardous Material Releases in Karst Terrains GEARY M. SCHINDEL* Karst Works, Inc. 11310 Whisper Dawn Street, San Antonio, Texas 78230
RUDOLPH ROSEN Professional Nonprofit Management Group LLC, 8501 Cocobolo Drive, Denton, TX 76207
GRAHAM M. SCHINDEL Arcadis, U.S. Inc., 410 N. 44th Street, Suite 1000, Phoenix, AZ 85008
Key Terms: Hazardous Materials, Karst, Source Water Protection, Emergency Response ABSTRACT In the United States, as much as 20 percent of the land surface is considered karst, and 40 percent of the groundwater for municipal and domestic consumption comes from karst aquifers. The U.S. Environmental Protection Agency has recognized karst aquifers as the groundwater type most vulnerable to contamination from anthropogenic sources, including hazardous materials and other pollutants. This paper presents a series of recommendations to assist first responders, water resource managers, and community leaders during a hazardous materials release in karst terrains. Hazardous materials releases have the potential to adversely impact human health and the environment. Karst aquifers are noted for rapid and direct recharge through caves, sinkholes, and sinking streams; groundwater velocities can exceed 1 km per day and are often complex and poorly understood. Best management practices in karst include pre-event planning, response, remediation, and long-term monitoring for hazardous materials releases. INTRODUCTION Hazardous materials (HAZMAT) occur and are widely used throughout industrialized society. HAZMAT may be released into the environment accidentally or intentionally. They are present at many retail and industrial sites, warehouses, home and garden centers, and along product pipelines, roadways, and railroad transportation corridors. When released to the environment, HAZMAT can impact groundwater and surface water resources, pose acute and *Corresponding author email: gearyschindel@karstworks.com
chronic threats to public health and safety, and threaten ecological systems and endangered species. The release of HAZMAT can occur as a result of natural and humancaused disasters, including floods, high winds, lightning, wildfires, accidental and intentional spills, earthquakes, vandalism, and terrorist activities. Rapid groundwater flow through karst aquifers can quickly carry HAZMAT as well as firefighting products to water wells and springs during emergency responses (Schindel and Rosen, 2021). In addition, some HAZMAT can volatilize and form explosive or toxic vapors that can migrate to the surface and into buildings or sewer lines (Quinlan et al., 1991). HAZMAT can quickly degrade public and private water supply wells, municipal and industrial water supplies, and surface waters (Johnson et al., 2010; Schindel, 2017, 2018; and Schindel and Rosen, 2021). Groundwater flow paths and velocities are poorly understood, site specific, and difficult to determine in karst aquifers (Schindel et al., 2004a). Once HAZMAT enter the aquifer, they can move quickly over large distances, becoming difficult and expensive to remediate. Proper pre-planning, event response, monitoring, and remediation efforts must be tailored to the unique characteristics of each karst aquifer. While there may be individuals with considerable experience in some local areas where groundwater flow paths in the aquifer have been mapped using tracer testing or other hydrogeologic tools, few comprehensive plans or programs have been developed that are detailed enough to address the unique nature of most karst watersheds and the complex geology of their underground aquifer systems (Schindel, 1986, 2019a). KARST In the United States, 20 percent of the land surface is karst, and 40 percent of the groundwater used for human
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 183–190
183
Schindel, Rosen, and Schindel
Figure 1. U.S. Geological Survey Karst Map of the Conterminous United States, 2020 (U.S. Geological Survey, 2020).
consumption comes from karst aquifers (Figure 1). The U.S. Environmental Protection Agency has recognized karst aquifers as the groundwater type most vulnerable to hazardous contaminants and pollution (Schindel et al., 1996; U.S. Environmental Protection Agency, 2002). Karst aquifers are unique because of their direct opening to the land surface, including enlarged fractures, sinkholes, caves, and sinking streams. Ford and Williams (1989) defined karst as a “terrain with distinctive hydrology and landforms arising from a combination of high rock solubility and well-developed secondary and tertiary porosity.” Huntoon (1995) defined karst as “an integrated mass-transfer system in soluble rocks with a permeability structure dominated by conduits dissolved from the rock and organized to facilitate the circulation of fluid.” Both of these definitions are important to understanding the unique properties and problems of working in karst. Karst watersheds are typically underlain by limestone or other highly soluble rocks, such as dolostone and gypsum. These rock types are partially dissolved over geologic time by chemical reactions and physical interactions with water (White and White, 1989). The dissolution process creates interconnected openings in the rock, thereby increasing its porosity and permeability. Over time, these openings can become conduits (1 to 2 cm in size up to about 0.5 m in diameter) and enlarge into caves (voids large enough for a person to enter). Many caves have conduits that can be measured in meters to tens of meters in diameter and have mapped lengths of over 1 km. Caves and conduits become integrated into underground networks with water discharging at springs. There are noted exceptional terrains that, 184
given sufficient geologic time, can also develop karst properties because all rocks, to some degree, are soluble in water. Noted examples include some sandstones (Shade, 2002) and even igneous rocks (Breisch, 1986). Karst terrains (which include both surface and subsurface features) may be identified by the presence of sinkholes, sinking streams, caves, interconnected voids, subsurface streams, enlarged fractures and faults, and springs that discharge water to the surface. However, many karst features can be difficult to detect. For example, sinkholes in the Edwards Aquifer region of south-central Texas may be broad and shallow, buried beneath fill (natural and anthropogenic), obscured by vegetation, covered by humanmade structures, or simply hard to find by inexperienced observers (Schindel, 2019a). Caves and conduits present beneath the soil and bedrock are commonly not detected by observation at the surface because they have no or very limited surface expressions. Light detection and ranging (LiDAR) is becoming an increasingly useful tool in helping to identify subtle karst features, especially in heavily vegetated terrain. Occasionally, caves are intersected by borings, road cuts, trenching, or during excavation for buildings, but such karst evidence is rarely accessible to HAZMAT emergency responders. Most caves and conduits have not been identified and recorded for many karst watersheds (Rosen et al., 2020). The absence of known karst features at a HAZMAT release site is not evidence that the site is not karst (Quinlan et al., 1995). Karst groundwater systems are commonly the source for springs that form the headwaters of many streams and rivers. Karst springs can range in discharge from a few liters per minute to as much as 5 to 10 m3/s during
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 183–190
Hazardous Material Releases in Karst Terrains
baseflow conditions. In addition, discharge can change in a matter of minutes or hours in response to rainfall and/or snowmelt events. Openings at the surface allow direct exchange of surface water with the subsurface and groundwater, providing little or no filtration or biological treatment of potential contaminants in the water (Schindel et al., 1996). Water can move quickly vertically and horizontally in karst aquifers with velocities commonly measured at over 1,000 m per day using groundwater tracers (Quinlan and Ewers, 1985; Worthington, 1999). Karst watersheds can range from a few hectares to tens of thousands of hectares and have flow paths extending many kilometers. These conditions allow for the rapid transport and spread of contaminants. Contaminants released into the aquifer can also imperil sources of public and private water supplies, headwaters to rivers, and sensitive cave and aquatic ecosystems that are home to endemic and endangered species.
their solubility is exceeded. Thus, DNAPLs are heavier than water and will form a free phase that will sink and enter fractures and sediments in the bottom of a conduit. They can be redistributed by turbulent flow in the groundwater, where they may re-solubilize, and/or they may redisperse and re-enter the water column. Common examples of DNAPLs are polychlorinated biphenyls and perchloroethylene. Insoluble contaminants may remain sequestered in soil or groundwater or become suspended during high-flow events. They may remain a source of low-level contamination over extended periods of time. Sequestered contaminants in karst aquifers can be difficult and expensive to investigate and remediate. Aquifer remediation may be technically infeasible and result in either expensive treatment systems at the point of withdrawal or abandonment of a water supply. Spring discharge may also be extremely difficult to retain and treat because of physical limitations at the spring as well as wide variations in discharge resulting from storm events.
THE NATURE OF HAZARDOUS MATERIALS
Pathways for Contamination
Hazardous solids, liquids, and gases can pose a shortor long-term risk to both surface-water and groundwater quality and can impact public water supply systems and ecological systems and endangered species. HAZMAT can be soluble or insoluble in water. Soluble HAZMAT can dissolve and be transported as a dissolved phase in water. Soluble HAZMAT may exceed drinking water standards as defined by the Safe Drinking Water Act (1974) and the Federal Water Pollution Control Act (1972). When the solubility limit of a compound is exceeded, it will have a higher or lower density than water and can form a non-aqueous dispersed phase liquid. HAZMAT that have a lower density than water form light non-aqueous phase liquids (LNAPLs) that float on the water surface. They can also volatilize (release vapors), becoming a gas that can migrate through fractures and conduits into sewer lines and buildings through crawlspaces and cracks in foundations (Stroud et al., 1986). Gasoline is a common example of an LNAPL that can migrate into a sewer or building and create an explosive condition if exposed to an ignition source (Stroud et al., 1986; Quinlan et al., 1991). Some gases, such as propane, are heavier than air and if released can enter depressions and caves and create explosive environments (Stroud et al., 1986). The movement of an LNAPL through an aquifer may be impeded where the cave ceiling extends below the water surface and acts as an “oil/water separator” (Quinlan and Ewers, 1985). Rapid groundwater flows may entrain LNAPLs in the water column through turbulence. Compounds that have a higher density than water can form dense non-aqueous phase liquids (DNAPLs) once
The many caves, sinkholes, fractures, faults, and other sensitive features characteristic of karst terrains provide direct pathways and allow rapid entry of material into groundwater. In many cases, karst features remain hidden, buried near the surface under shallow soils, and have not been identified. Sensitive features may only become apparent on-site during emergency response or when spill remediation efforts are under way. In addition, many recharge features are located within drainage ways and may be obscured/buried by gravel, cobbles, and sediments. Buried features are very difficult to detect and leave little indication of their presence on the surface. However, they can rapidly receive and convey water and other liquids to an aquifer (Quinlan and Ewers, 1985). Thick soils in some karst areas may also retain residual HAZMAT and release the contaminant(s) slowly over time or rapidly during storm events (Quinlan and Ewers, 1985; Stroud et al., 1986). Water wells can also be a pathway for contaminants to enter an aquifer. Water wells include both public water supply wells and privately owned wells that are used for domestic, municipal, industrial, livestock, and agricultural purposes. Well construction over the last 100 years has included a wide range of drilling and construction methods. Wells may range from those that have been fully cased and grouted to isolate groundwater from surface activities and other impacts to wells that contain no casing or grout and remain open to the surface. The top of the well casing may be at or below land surface level (or below flood levels), which can allow surface water and other materials to readily enter the well and aquifer. Poorly constructed or maintained wells
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 183–190
185
Schindel, Rosen, and Schindel
in active use or those that have been abandoned or that are unsecured can become conduits for surface water or shallow aquifers containing contaminants to rapidly enter groundwater (Green et al., 2006). In some areas, injection wells have been drilled in the bottom of sinkholes and are open to the surface to facilitate drainage of stormwater. In the United States, these are defined as Class V injection wells and are regulated under federal- and, in some cases, state-administered underground injection control programs. Release of contaminated surface water in a sinkhole or drainage way containing an injection well may result in direct contamination of the aquifer. Some stormwater retention systems and modified sinkholes that allow infiltration into the subsurface may also be considered Class V injection wells and can also facilitate groundwater contamination (Stroud et al., 1986). Manufactured features such as trenches, pipelines, storm and sanitary sewers, drainage systems, water wells, and building foundations can also collect and act to direct water and contaminants into the subsurface. Water-Quality Impacts The widespread creation, storage, transport, and use of HAZMAT in the United States pose a risk to the environment, including highly vulnerable karst aquifers. Most public water treatment systems utilizing groundwater are only required to use chlorination to supply safe drinking water. Private (domestic and agricultural) water systems commonly have no water-quality monitoring or treatment at all. Hazardous materials can rapidly enter a groundwater system. Within hours or days, this can result in the exceedance of Safe Drinking Water Act and Clean Water Act standards, which can then result in acute and chronic human health impacts and ecological impacts. Treatment or replacement of a contaminated public water supply can cost millions of dollars and require months to years to complete. Treatment facility costs are in addition to the long-term cost of groundwater remediation, where remediation is even possible. Groundwater contamination can also impact subsurface and surface ecosystems, which can result in impacts to related ecosystem services including fisheries and recreational opportunities. Given the potentially high direct and indirect costs of even a single catastrophic release of HAZMAT, resource managers and first responders need education materials, training, and tools to protect aquifer and surface-water quality during emergency response activities. BEST MANAGEMENT PRACTICES Best management practices (BMPs) are methods that are considered to be effective and practical ways to prevent or reduce non-point and point pollution and to help achieve 186
water-quality goals. BMPs include measures to prevent and mitigate pollution and the potential impact to human and environmental receptors. Generally, BMPs are designed to protect water quality. This can also protect human health and safety, minimize impacts to the local economy and human environment, maintain aquatic ecosystems, and protect threatened and endangered species. The BMPs we describe are intended for implementation by state, county, city, municipal, and government employees, environmental contractors, and others responsible for responding to, managing, or regulating spills and releases of HAZMAT. The BMPs can be applied in an aquifer’s contributing and recharge zones and thus will provide new and additional guidance for responding to the release of HAZMAT in karst watersheds. Implementation of BMPs for planning and emergency response will enable emergency responders to improve the resilience of karst aquifers through prevention, reduction, and mitigation of disaster and disaster-related impacts to drinking water supplies, public health, environmental quality, and the economy. BMPs provide new options for emergency response and include added science-based planning, standards, recommended advanced planning, training, specific curricula for first responders, and disaster response options that focus on accumulated data, predictive modeling, and on-site, real-time evaluation of the transport and fate of HAZMAT releases to karst aquifers (Schindel and Rosen, 2021). Educational materials and training based on BMPs that conform to Hazardous Materials Standards for Responders (National Fire Protection Association, 2022) have been developed through the Texas A&M Engineering Extension Service (Emergency Services Training Institute, 2021; Schindel and Rosen, 2021). This training was made available out of concern for potential contamination of the karstic Edwards Aquifer during emergency response through the City of San Antonio with funding from the voter-approved Proposition 1 Edwards Aquifer Protection Venue Project. The BMPs we recommend fall into three categories commonly used in emergency response training and planning materials: (1) pre-event planning, (2) event response, and (3) post-event response. Table 1 provides an outline of BMP methods. Pre-Event BMPs Pre-planning for potential releases will help responders and public officials minimize impacts from HAZMAT spills. While predicting when and where an emergency event might take place is difficult, planning can focus on locations with a history of spills where HAZMAT are stored or may be highly susceptible to potential HAZMAT release. Such locations include dangerous intersections, steep grades, stream crossings in transportation corridors, and HAZMAT manufacturing and storage locations.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 183–190
Hazardous Material Releases in Karst Terrains Table 1. Summary list of BMPs. Pre-Event BMPs Develop a comprehensive GIS mapping program Perform a baseline hydrogeologic study Perform tracer testing Locate potential HAZMAT sources Assess HAZMAT sites Conduct water-quality sampling Identify and contract with hydrologic consultant(s) Develop public communication plans Event Response BMPs Identify the time of release, location, material(s), and estimated volumes Retain, remove, and remediate any HAZMAT Apply location-specific model(s) Deploy non-toxic fluorescent dyes Implement the water-quality sampling program Notify water users who potential may receive contaminated water Post-Event BMPs Continue to implement the water-quality sampling program Continue to sample for groundwater tracers where used Develop remedial actions Inform potential water users Evaluate and remediate any materials
Development of BMPs for use locally can be done by expert stakeholders working together to draft and review BMPs for karst aquifer protection (Schindel and Rosen, 2021). The stakeholder group should be formed under a government or municipal utility entity. Pre-planning for HAZMAT incidents should include the following efforts. • Develop a comprehensive geographic information system (GIS) mapping program for the defined karst watershed. Acquire relevant data layers to include the following: (1) topography and LiDAR data; (2) surface geology, soils, and vegetation coverage; (3) stormwater containment structures and assigned characteristics; (4) surface runoff flow paths including culverts, ravines, and creek and stream beds; (5) locations and nature of karst features, such as creek beds, sinkholes, caves, springs, faults, and fractures; (6) footprints of structures containing HAZMAT; and (7) and locations of domestic and public water wells, surface-water supply intakes, and similar structures. • Perform a baseline hydrogeologic study to determine the potential boundaries of the contributing zone and recharge zone of the groundwater basin(s) of concern. This may include the mapping of geologic and hydrologic boundaries and potentiometric surfaces, topography, tracer testing, and water-quality data. • Perform tracer testing in the basin to help define boundaries of the aquifer and confirm groundwater flow directions and velocities under varying aquifer conditions (Schindel et al., 1996; Schindel, 2019b). • Locate potential HAZMAT sources by identifying manufactures, storage, use, and transport corridors of large
quantities of HAZMAT. These locations include large retail sites, underground storage tanks, industrial and agricultural sites, pipelines, and truck and rail transportation corridors. Incorporate perimetrical flow information from these locations into GIS map layers, including the surface coverage and stormwater containment structures that can potentially contain HAZMAT released during disaster-related events. • Assess HAZMAT sites to determine if secondary or tertiary containment and spill prevention, control, and countermeasure plans exist for the site and are up to date. • Conduct aquifer water-quality sampling at public water supply wells and springs to determine background concentrations of common HAZMAT present or used within the area. These data can be used for comparison with surface-water runoff during a HAZMAT incident. • Identify and contract with hydrogeologic consultant(s) to assist with or direct the response to a HAZMAT incident. The consultant may utilize the data generated through advanced planning to support work with first responders and others to initiate a water-quality sampling program, identify potential karst receptors, and make recommendations to protect these features. The consultant may recommend a tracer testing program as a surrogate for released material. Water-quality and tracer test data can be used to make resource decisions related to potential impacts to public and private water supplies and groundwater remediation. • Develop public communication plans, including appropriate language options, based on application of models, dye tracing, and HAZMAT transport and fate information. Event Response BMPs • Identify the time of release, location, material(s), and estimated volumes released as quickly as possible after an incident has been reported. • Retain, remove, and remediate any HAZMAT on site and on the ground surface. Take efforts to ensure that HAZMAT is not flushed with water or directly moved into the aquifer or drainage paths to the aquifer. • Apply location-specific model(s) for predicting surface-water flow and aquifer vulnerability including existing tracer testing data. Use available data to predict the range of concentrations that may be detected at receptors of concern. • Deploy non-toxic fluorescent dyes for use in tracing contaminant flow and fate if deemed appropriate by the hydrogeologic consultant. On-site non-toxic fluorescent dye tracing during disaster response can be used to collect data for planning and directing data-based responses to protect public health and reduce environmental impacts. Incorporate information as appropriate for on-the-ground
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 183–190
187
Schindel, Rosen, and Schindel
response. Tracing or contaminant sampling should be initiated within 24 hours to be most effective in identifying water wells, springs, and surface waters that may be impacted by the release. Fluorescent dyes are much quicker and cheaper to analyze than most HAZMAT constituents. • Implement the water-quality sampling program designed during pre-event planning to detect the material released. Modify the sample plan as required. • Notify water users who potentially may receive contaminated water, including managers responsible for springs and surface waters, public and private water wells, and aquatic ecosystems. Notifications are based on the hydrogeologic assessment prepared during preevent planning. Post-Event BMPs • Continue to implement the water-quality sampling program to evaluate the fate of HAZMAT released into the aquifer. Sampling should be performed at wells, springs, and other potential receptors and continued through at least one stormwater event and until background concentrations are reached. Sampling should be event (storm) driven and not part of commonly used quarterly monitoring programs. • Continue to sample for groundwater tracers, where used, until samples are undetectable and at least one stormwater event has passed. Additional tracer tests may be required to help refine understanding of local and regional groundwater systems and HAZMAT fate and transport. • Develop the remedial actions that are determined to be necessary based on results from models, dye tracing, and water sample collection and testing for HAZMAT. • Inform potential water users of the status of the HAZMAT response and findings and issue health advisories or other disaster-related communications as appropriate. • Evaluate and remediate any materials that have collected in stormwater retention basins. Ensure proper disposal of the HAZMAT.
RECOMMENDED ACTIONS TO SUPPORT BMPS Public safety should remain the utmost priority during emergency events, but with proper planning, acquisition of essential data, and efficient communication, many environmental concerns can also be addressed. Hydrogeologic evaluation of the karst watershed and adjacent watersheds of concern should be performed as part of the pre-event planning. These activities include the following: 188
• Use LiDAR to identify subtle karst features that may be obscure or difficult to observe because of thick vegetation, or where there are limitations on property access. LiDAR has emerged to be a powerful tool that can be used in many areas and should be included in the GIS mapping process. • Use cave locations and cave maps to aid in determining the extent of karst development in an area. However, the absence of cave data may not be reflective of the absence of karst. Some areas have not been evaluated or are not accessible to the caving community. In addition, most state cave surveys are managed by volunteers, and the data may not be readily available to researchers. If available, these types of data can prove to be very valuable. • Use geologic maps to define areas that are more or less likely to form karst and to help define areas for evaluation and inclusion in the study. • Use tracer testing to help define the direction of groundwater flow, groundwater velocities, and estimate contaminant concentrations to receptors of concern (Schindel et al., 2004b).
Data from evaluations described above can be used to develop science-based BMPs for on-site and realtime use during HAZMAT response. Of the evaluation methods, tracer testing is the most specialized and possibly the most useful tool in pre-planning activities and during a HAZMAT response. Tracer testing should be performed by experts with specialized training in dyetracing methods (Alexander and Quinlan, 1997). Dye tracing is a tool that can be rapidly deployed to determine the fate and transport of HAZMAT in karst aquifers in near real time (Schindel et al., 2003; Johnson et al., 2010). While dye tracing has limits when used in karst due to the complexity of karst environments, tracing studies have been used in them for more than 100 years to help characterize groundwater flow (White and White, 1989) and provide empirical data on fate and transport of constituents in the aquifer (Schindel et al., 2005). Results from tracer testing can be quickly obtained and provide greater cost-effective results than the complex analysis required for many contaminants. Waterquality analysis is required to determine matters of public health, such as if water discharged from a well or spring meets water-quality standards. Consultants and laboratories that will conduct tracer testing and water-quality analysis should be identified and have contracts in place as part of the pre-planning process. Depending on the volume of material released, flows that enter a drainage way may travel down gradient for thousands of feet (hundreds of meters) beyond the event/ property boundary. The following guidance and emergency
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 183–190
Hazardous Material Releases in Karst Terrains
event mitigation BMPs are recommended for implementation as quickly as possible: • Identify the leading edge of the contaminant flow(s) on the ground surface. • Identify the nearest down-gradient points of potential entry into the aquifer (sensitive features such as caves, sinkholes, fractures and faults, sinking streams, storm drains, stormwater retention basins, and active and abandoned water wells). • Identify the best method or combination of methods to control the HAZMAT flow, along with control of firefighting products, if relevant. • Attempt to determine the amount of material released, volume of water, including amounts of other substances added to control fires (if relevant), and the amount of material that has been recovered. This water, including HAZMAT and firefighting products, should be blocked or diverted from flowing into the identified storm drains, ravines, creeks, sinkholes, sinking streams, caves, fractures and faults, wells, or drainage ways. Whenever possible, use covers, berms, booms, sandbag dams, or plastic sheeting to prevent or minimize HAZMAT from reaching storm drains, drainage features, surface waters, sinkholes, caves, and other sensitive features or other pathways into the subsurface. Stormwater retention basins, where available, may be used for temporary storage of HAZMAT or firefighting products, with attention to compatibility and design of the retention basin. Identification and evaluation of stormwater retention basins during pre-planning should include the evaluation of the material lining the basin (if any) for compatibility with potential HAZMAT. If necessary, the retention basin liner may need to be replaced with other types of liners. Firefighting products should be retained until they can be properly disposed of or otherwise remediated.
CONCLUSION AND IMPLEMENTATION Contamination of karst aquifers can occur from natural causes, but contamination that threatens public health and aquatic environments is most commonly caused by human action or inaction (Rosen, 2014). The best means to secure karst groundwater supplies from becoming contaminated is to prevent HAZMAT from reaching entry points into the aquifer. This paper provides a set of BMPs for use by resource managers and first responders to protect the quality of water in karst aquifers from HAZMAT discharge. BMPs are provided for preplanning, response during an emergency, and cleanup and remediation after an event.
REFERENCES ALEXANDER, E. C., Jr., AND QUINLAN, J. F., 1997, Practical tracing of groundwater with emphasis on karst terranes. In Guidelines for Wellhead and Springhead Protection Area Delineation in Carbonate Rocks, Appendix B, 2nd ed.: U.S. Environmental Protection Agency, Region 4 (GWDW-15), EPA 904-B-97-003, May 1997, 155 p. BREISCH, R. L., 1986, Greenhorn caves, the granddaddy of granite caves: NSS News, Vol. 44, No. 4, p. 86. EMERGENCY SERVICES TRAINING INSTITUTE, 2021, Run-Off Control for Karst Environments: Participant Manual, v. 3.2.2021: Texas A&M Engineering Extension Service, Emergency Services Training Institute, College Station, TX, 97 p. FEDERAL WATER POLLUTION CONTROL ACT, 1972, 33 U.S.C. §§ 1251–1387. FORD, D. C. AND WILLIAMS, P. W., 1989, Karst Geomorphology and Hydrology: Unwin Hyman Inc., London, U.K., 601 p. GREEN, R. T.; PAINTER, S. L.; SUN, A.; AND WORTHINGTON, S. R. H., 2006, Groundwater contamination in karst terrains: Water, Air, and Soil Pollution: Focus, Vol. 6, No. 1–2, pp. 157–170. https://doi .org/10.1007/s11267-005-9004-3. HUNTOON, P. W., 1995, Is it appropriate to apply porous media groundwater circulation models to karstic aquifers? In El-Kadi, A. I. (Editor), Groundwater Models for Resources Analysis and Management: Lewis Publishers, Boca Raton, FL, pp. 339–358. JOHNSON, S. B.; SCHINDEL, G. M.; AND VENI, G., 2010, Tracing Groundwater Flow Paths in the Edwards Aquifer Recharge Zone, Panther Springs Creek Basin, Northern Bexar County, Texas: Report No. 10-01, Edwards Aquifer Authority, San Antonio, TX, 112 p. https:// www.edwardsaquifer.org/wp-content/uploads/2019/05/2010_ Johnson-etal_PantherSpringsCreekBasinFlowpaths.pdf NATIONAL FIRE PROTECTION ASSOCIATION, 2022, NFPA 470, Hazardous Materials/Weapons of Mass Destruction (WMD) Standard for Responders: National Fire Protection Association, Boston, MA. https://www.nfpa.org/codes-and-standards/all-codes-and-standards/ list-of-codes-and-standards/detail?code¼470 QUINLAN, J. F. AND EWERS, R. O., 1985, Groundwater flow in limestone terranes: Strategy rationale, and procedures for reliable, efficient monitoring of ground water quality in karst areas. In Proceedings of the 5th National Symposium and Exposition on Aquifer Restoration and Ground Water Monitoring, Columbus, Ohio: National Water Well Association, Worthington, OH, pp. 197–234. QUINLAN, J. F.; RAY, J. A.; AND SCHINDEL, G. M., 1995, Intrinsic limitations of standard criteria and methods for delineation of groundwater source protection areas (springhead and wellhead protection areas) in carbonate terranes: Critical review, technically-sound resolution of limitations, and case study in a Kentucky karst. In Beck, B. F. (Editor), Karst Geohazards—Engineering and Environmental Problems in Karst Terranes: A. A. Balkema, Rotterdam, Netherlands, pp. 525–537. QUINLAN, J. F.; SCHINDEL, G. M.; AND LYONS, J. K., 1991, It ruins my day when a cave explodes: Risks associated with investigating hazardous environments in karst terranes, and protocol for safely doing so. In Abstracts with Programs of the Geological Society of America, 1991 Annual Meeting of the Geological Society of America; October 22; San Diego (California): Geological Society of America, Boulder, CO, Vol. 23, No. 5, p. 325. ROSEN, R. A., 2014, Texas Aquatic Science: Texas A&M University Press, College Station, TX, 200 p. ROSEN, R. A.; HERMITTE, S. M.; MACE, R.; AND WADE, R., 2020, Internet of Texas water: Use cases for flood, drought, and surface water–groundwater interactions: Texas Water Journal, Vol. 11, No. 1, pp. 133–151. https://doi.org/10.21423/twj.v11i1.7109. SAFE DRINKING WATER ACT, 1974, Title XIV of Public Health Service Act. Pub. L. No. 93-523, 88 Stat. 1660 (Dec. 16, 1974).
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 183–190
189
Schindel, Rosen, and Schindel SCHINDEL, G. M., 1986, Aquifer protection planning for critical karst ground water supplies: A coordinated local and state approach. In Proceedings of the Environmental Problems in Karst Terranes and Their Solutions Conference, October 28–30, 1986, Bowling Green, Kentucky: National Water Well Association, Columbus, OH, pp. 425–435. SCHINDEL, G. M., 2017, Source water protection strategies for karst aquifers. Abstract: Sustainable management of groundwater resources. In Posavec, K. and Markovi c, T. (Editors), 44th Annual Congress of the International Association of Hydrogeologists: Croatian National Chapter of the International Association of Hydrogeologists, Dubrovnik, Croatia, p. 50, September 25–29, 2017. SCHINDEL, G. M., 2018, Recommended strategies for the response to hazardous materials releases in karst. In White, W. B.; Herman, J.; Herman, E.; and Rutigliano, M. (Editors), Karst Groundwater Contamination and Public Health: Beyond Case Studies: Springer Nature, Cham, Switzerland, pp. 255–260. SCHINDEL, G. M., 2019a, Genesis of the Edwards (Balcones Fault Zone) Aquifer. In Sharp, J. M.; Green, R. T.; and Schindel, G. M. (Editors), The Edwards Aquifer: The Past, Present, and Future of a Vital Water Resource: Memoir 215, Geological Society of America, Boulder, CO, pp. 9–18. https://doi.org/10.1130/2019.1215(02). SCHINDEL, G. M., 2019b, Recommended Best-Management Practices for Response to Spills in the Edwards Aquifer of South-Central Texas: Project Report, Edwards Aquifer Water Quality Protection from Catastrophic and Low to Mid-level Effects of Discharge of Hazardous and Polluting Materials from Contaminated Water Runoff during Emergency Response, Institute for Water Resources Science and Technology, Texas A&M University San Antonio, San Antonio, TX. https://libguides.tamusa.edu/ld.php?content_id¼57497570 SCHINDEL, G. M.; HOYT, J. R.; AND JOHNSON, S. B., 2004a, Edwards Aquifer, United States. In Gunn, J. (Editor), Encyclopedia of Cave and Karst Science: Fitzroy Dearborn Publishers, London, pp. 648–651. SCHINDEL, G. M.; JOHNSON, S. B.; ALEXANDER, E. C., Jr.; WORTHINGTON, S. R. H.; AND DAVIES, G. J., 2004b, Quantitative tracing as a predictive tool to assess the potential impacts of hazardous materials to water supplies and environmental receptors: Geological Society of America Abstracts with Programs, Vol. 36, No. 5, p. 396. https:// gsa.confex.com/gsa/2004AM/webprogram/Paper79163.html SCHINDEL, G. M.; JOHNSON, S. B.; AND VENI, G., 2005, Tracer tests in the Edwards Aquifer recharge zone: Geological Society of America Abstracts with Programs, Vol. 37, No. 7, p. 216. https://gsa .confex.com/gsa/2005AM/webprogram/Paper95392.html
190
SCHINDEL, G. M.; QUINLAN, J. F.; DAVIES, G. J.; AND RAY, J. A., 1996, Guidelines for Wellhead and Springhead Protection Area Delineation in Carbonate Rocks: U.S. Environmental Protection Agency, Region IV Groundwater Protection Branch, Atlanta, GA, 195 p. https://nepis .epa.gov/Exe/ZyPDF.cgi/9101WWCC.PDF?Dockey¼9101WWCC .PDF SCHINDEL, G. M. AND ROSEN, R. A., 2021, Best management practices for fire fighting in the karstic Edwards (Balcones Fault Zone) Aquifer of south-central Texas: Texas Water Journal, Vol. 12, No. 1, pp. 1–9. https://doi.org/10.21423/twj.v12i1.7110. SCHINDEL, G. M.; WORTHINGTON, S. R. H.; DAVIES, G. J.; ALEXANDER, E. C., Jr.; AND JOHNSON, S. B., 2003, Quantitative tracers as contaminant surrogates: An important tool for planning and managing source water protection areas: Geological Society of America Abstracts with Programs, Vol. 34, No. 7, p. 281. https://gsa.confex .com/gsa/2003AM/webprogram/Paper67716.html SHADE, B. L., 2002, The Genesis and Hydrogeology of a Sandstone Karst in Pine County, Minnesota: Master’s Thesis, University of Minnesota, Department of Earth & Environmental Sciences, 171 p. STROUD, F. B.; CRAWFORD, N. C.; AND RIGATTI, M. M., 1986, U.S. Environmental Protection Agency emergency response to toxic fumes and contaminated groundwater in karst topography: Bowling Green, Kentucky. In Proceedings of the Environmental Problems in Karst Terranes and Their Solutions Conference, October 28–30, 1986: National Ground Water Association, Bowling Green, KY, pp. 197–225. U.S. ENVIRONMENTAL PROTECTION AGENCY, 2002, Delineation of Source-Water Protection Areas in Karst Aquifers of the Ridge and Valley and Appalachian Plateaus Physiographic Provinces: Rules of Thumb for Estimating the Capture Zones of Springs and Wells: EPA 816-R-02-015, U.S. Environmental Protection Agency, Washington, D.C., 41 p. U.S. GEOLOGICAL SURVEY, 2020, Karst Map of the Conterminous United States—2020: U.S. Geological Survey. https://doi.org/10 .5038/9781733375313.1003. WHITE, W. B. AND WHITE, E. L., 1989, Karst Hydrology, Concepts from the Mammoth Cave Area: Pennsylvania State University/Van Nostrand Reinhold, New York, NY, 346 p. WORTHINGTON, S. R. H., 1999, A comprehensive strategy for understanding flow in carbonate aquifers. In Palmer, A. N.; Palmer, M. V.; and Sasowsky, I. D. (Editors), Karst Modelling: Karst Waters Institute, Charlestown, WV, pp. 30–37.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 183–190
Economic Exclusion and Forgotten Floodplains on Karst Terrain SARAH A. BURGESS Tacoma, WA 98444
LEE J. FLOREA* Washington Geological Survey, Olympia, WA 98501
Key Terms: Lost River, Orangeville Rise, Hydrography, Flood Insurance, Mitchell Plateau, Environmental Justice ABSTRACT Flood risk models in karst landscapes often use methods that do not capture the complex linkages between surface water and groundwater flow. Therefore, published maps for access to the National Flood Insurance Program may exclude those who experience regular inundation. Using simple geographic information system–based inundation models, we demonstrated how one town in Indiana, Orleans, in the classic karst landscape of the Lost River watershed, has been systematically excluded from flood insurance risk maps because Federal Emergency Management Agency–produced models do not consider the town to be prone to floods, despite well-documented evidence to the contrary. Our simple models connect missing elements to the National Hydrography Dataset and show that published flood insurance rate maps significantly underpredict the possible scope of floods in this watershed. Recently proposed highway infrastructure included alternatives that would have bisected the Lost River watershed and could have potentially increased flooding concerns. Unless comprehensive studies detail the actual risk of floods as (1) part of the natural landscape response in karst, (2) a natural consequence of more extreme events as climate changes, and (3) a consequence of impaired flow routes possible from road construction, towns such as Orleans may experience amplified economic exclusion. INTRODUCTION On May 1–3, 2010, torrential rainfalls caused historic flooding in western Kentucky, Tennessee, and northern Mississippi (National Weather Service, 2010). News outlets carried stories and images from Nashville, TN, as rising levels of the Cumberland River covered downtown *Corresponding author email: Lee.Florea@dnr.wa.gov
and reached 1,000-year-flood records. Damage was estimated at $2.3 billion, and 26 people died (NWS, 2010). It is likely that damage would have been considerably worse if repairs on levees in the Nashville metropolitan area had not been completed or if water levels in Lake Cumberland in Kentucky had not been lowered by 45 ft (13.7 m) to conduct repairs on Wolf Creek Dam to mitigate a second round of embankment leakage discovered since the 1941–1951 construction. The repairs at Wolf Creek Dam installed a second concrete diaphragm into competent bedrock underlying large karst features uncovered during original construction. Across the country, the Swift Canal Project, part of a hydroelectric dam completed in 1957 in southwest Washington State, experienced a catastrophic failure on April 21, 2002 (Regan and Langnion, 2008). This failure was the most dramatic of a series of intrusions of impounded water into sinkholes and through lava tubes in the underlying lava flow on the south slopes of Mount Saint Helens. During excavations and repairs in 1958 and again in 1973, more than 50 sinkholes were discovered each time and filled. Despite these repairs, water seeping below the basalt excavated a 1–2 acre (4047– 8094 m2) void beneath the foundation of the dam. Due to the buried nature of the karst, no detection was possible until the drainage intersected cracks at the toe of the embankment, making intervention impossible. The outcome was a massive blowout along the canal embankment that flooded the highway and damaged a nearby power station (Regan and Langnion, 2008). Wolf Creek Dam and Swift Reservoir are about 3,860 km apart geographically, are underlain by vastly different geologies (karst-forming carbonates versus volcanic strata), and had different outcomes (mitigation versus failure). Nonetheless, the two incidents are related by commonalities in physical hydrology and historical context. Impounded water enlarged conduits beneath and undermined both structures, which were built in the same time period of U.S. infrastructure construction. Comparing these incidents provides a useful frame of reference for the inconsistencies in planning for and management of karst-specific geologic hazards.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
191
Burgess and Florea
Another lesson from these events is the tenuous nature by which the United States has developed infrastructure. Major cities, roadways, and housing are often located in flood-prone areas, and when unexpected weather events strike, steps are taken to further engineer mitigation strategies to maintain this infrastructure. As climate models foretell greater instability with an increase in extreme events in the Midwestern United States (Pryor et al., 2014), we can expect amplified concerns about the cumulative cost of maintaining aging infrastructure and the need for new construction in areas prone to floods. One part of this cost is the National Flood Insurance Program (NFIP), which has provided an economic infrastructure to share the risk of loss due to flooding and to regulate development in existing floodplains since 1968. The Federal Emergency Management Agency (FEMA) defines a floodplain as “. . .lowland and relatively flat areas adjoining inland and coastal waters including, at a minimum, that area subject to a one percent or greater chance of flooding in any given year” (CFR, 2018, Title 44). The standard methods accepted by FEMA for floodplain mapping synthesize topography, stream gauging, channel morphology, and meteorology to predict the likelihood of an area to flood in a given year (FEMA, 2019). However, these models commonly under-represent the scope of flooding and the risk to infrastructure. First Street Foundation (2021), for example, found that 25 percent of critical infrastructure and 23 percent of all road segments in the United States are at risk of becoming inoperable or impassable, respectively. In another recent analysis (Kim, 2023), the Foundation for Appalachian Kentucky found that nearly 80 percent of all homes inundated during catastrophic 2022 floods in eastern Kentucky were not included in FEMA-designated high-risk areas. This study explored one subset of at-risk infrastructure, highways, in karst-dominated watersheds in the United States, where FEMA-utilized methodologies have reduced utility or are never employed. These methods are specifically less useful because intermittently active surface channels are under-drained by subterranean conduits that convey infiltration from the epikarst, sinkholes, and sinking streams to springs where the groundwater again becomes surface water. In other words, surface stream networks are often disconnected, if present at all. Scientific understanding of karst has greatly evolved; however, consideration of karst in watershed planning and regulation by local policy makers is generally not comprehensive, is inconsistently considered by state agencies, and is absent at the federal level despite repeated calls for change (Kemmerly, 1981). Ignoring the unique hydrology of karst when mapping flood hazards excludes communities in karst landscapes from insurance-based economic protections available elsewhere. Rural communities living on karst have few 192
resources to mitigate unprotected flood loss and damage. We explored a case study from the rural community of Orleans, Indiana, in the Lost River watershed, where the Indiana Department of Transportation (INDOT) recently considered a new highway construction project across the core of the karst of south-central Indiana—the Mid-States Corridor project (Lochmueller Group, 2022). Despite more than 100 years of documented karst flooding, the catchment that includes Orleans is completely excluded from NFIP consideration (Malott, 1952; Bayless et al., 2014) (Figure 1). This documentation and history of karst science in the vicinity of Orleans, combined with the Mid-States Corridor project proposal, make it a timely and impactful case study. We created geographic information system (GIS)–based inundation models that combined available elevation data with hydrography and stream gauging. We concluded that Orleans and similar communities have grossly underestimated flood risk because karst hydrology and local knowledge are not broadly considered. Karst Flooding In simplest terms, floods occur where rates of water input exceed water throughput. In large riverways, rising waters overtop natural levees, spill into adjacent floodplains, and leave observable scars that aggregate into the record of channel migration and sedimentation. Karst landscapes, in contrast, can flood quite differently. Underground conduits transmit water in a fashion similar to engineered storm drains, where the maximum discharge is proportional to the hydraulic radius and the hydraulic head and inversely proportional to the frictional resistance—described by the Darcy-Weisbach equation (Brown, 2003). If the recharge to the conduit is greater than the discharge capacity, water will pond in the contributing sinkholes and sinking streams. Since surface streams that carry this excess water are less frequent in karst, the water levels will rise to balance recharge with the discharge from the system—in essence, the sinkholes and sinking streams are the karst floodplain and commonly form a discontinuous surface-water network that present challenges in hydraulic modeling methods, which typically require contiguous waterways. During the May 2010 event that opened this paper, the city of Bowling Green, KY, and the greater southcentral Kentucky karst landscape experienced historic rainfall and region-wide flooding—the second author lived in Bowling Green at the time of this event. Initially, runoff collected in sinkholes and sinking streams, outpacing the carrying capacity of the caves. In an average storm event, water levels recede in the sinkholes and caves soon after the precipitation has ended and direct runoff ceases; the rate that ponded water drains is set by the size of the sinkhole throat and nature of the connection to the underlying caves.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
Flood Risk and Economic Exclusion
Figure 1. Photographs of flooding of roads and fields in the Lost River watershed on May 3, 2011. Left column: water on roads in the town of Orleans; middle column: flooding of roads and fields by the dry bed of the Lost River; right column: flooding of Orangeville Rise, a terminal spring for the karst in the Lost River watershed. Flooding at terminal springs causes backflooding in dry beds and widespread flooding in karst terrain. Photos by the Indiana Geological and Water Survey.
In the events of May 2010, however, a second complicating factor arose, and it was particularly noticed in Bowling Green. Water levels continued to rise following the end of the precipitation event, and residents noticed that the rise in water level originated from below the ground surface. The second author personally documented boils of water rising out of sinkholes, stormwater injection wells, and manhole covers. While not properly documented or monitored at the time, these events are consistent with the concept of a “valley tide” (Dougherty, 1983; Simpson and Florea, 2009), where an increased pressure head in the underlying caves causes water levels to rise in artesian conditions to a potentiometric surface above ground level in places where direct connectivity exists. More generally, valley tides are the surface-observable manifestation of the epiphreatic zone of karst aquifers (Gabrovšek et al., 2018), where water levels regularly rise and fall in accordance with singular events or seasonal changes in rainfall. Fluctuations in the epiphreatic surface in a karst landscape are a transient signal that originates in one of two possible ways. The first is the migration of a pulse through the system (Baedke and Krothe, 2001) when sinkholes and sinking streams in headwaters receive considerably more recharge than the downstream portions closer to the spring, such as allogenic-dominated systems in the Cumberland Plateau of Kentucky and Tennessee (Simpson and Florea, 2009) or the classic karst landscapes of Slovenia and Croatia (Pekaš et al., 2012). The second mechanism for valley tides is ponding or reversed flow, called “backflooding,” resulting from a
delayed flood pulse in the riverway to which the karst system drains (Albéric, 2004). Examples include spring reversals along the Suwannee, Santa Fe, and Withlacoochee Rivers of west-central Florida (Gulley et al., 2011) and at Mammoth Cave National Park in Kentucky, 40 km northeast of Bowling Green (Kipper, 2019). The May 2010 event in Bowling Green combined both valley tides and backflooding. The 142 km2 surface catchment for Lost River Cave (not to be confused with the Lost River in Indiana) south and east of the city routed a major storm pulse from the headwaters 19 km distant (Crawford, 2005) through two intermittent karst lakes, each over 1.5 km in diameter, through an underground cave network to the Lost River Blue Hole located in the Lost River Karst Window, and into the mouth of Lost River Cave. From there, the storm pulse traveled for over 2 km in a large cave passage under the southwest portion of Bowling Green to the Lost River Rise. In addition, the Barren River, to which Lost River Cave drains, rose for a lengthy period following the rainfall event. These river waters backflooded Jennings Creek upstream for 1.6 km to the Lost River Rise. The combined effects, observed by the second author, were ponded water in sinkholes for several days after the event concluded. In the most extreme instances, flood waters rose to 5 m above the land surface in some sinkholes, or more than 15 m above base flow as measured using topographic data. This example from Bowling Green highlights a few important messages about the complexity of karst landscape hydrogeology: (1) Floodplains in karst are regularly disconnected from major surface waterways and
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
193
Burgess and Florea
Figure 2. Map of the western portion of the Lost River karst basin, modified from Florea et al. (2018). Dye traces (blue lines) connect sites of dye injection (open triangles) to spring where dye was recovered (white circles). Additional sinking streams are shown as solid triangles. The 0.61km-wide Mid-States Corridor Alternative Route O is overlain in purple (with an adjacent 1.61-km-wide buffer from the center line in light purple).
include sinkholes, sinking streams, and intermittent lakes; and (2) the hydroperiod of these floodplain wetlands in karst can be guided by both landscape runoff and the potentiometric surface of groundwater in the underlying conduit system. Establishment and maintenance of critical highway infrastructure in karst must therefore be cognizant of both above points because highway construction can significantly alter runoff characteristics and alter the geometry and areal extent of these karst wetlands. As a consequence, the communities, farms, and environment may experience regional impacts, including more intense flooding, not limited to the right-of-way or the adjoining buffer usually considered as part of an environmental impact statement (EIS) as required by state and federal environmental law (Florea et al., 2002). As FEMA-based models often exclude these potentially flood-prone areas 194
in karst, impacted residents may not have access to NFIP compensation for loss. STUDY AREA Orleans is an incorporated town of about 3,500 people in Orange County, IN (Figure 2); it lies entirely in the Lost River watershed, a 948 km2 area of the Crawford Uplands and Mitchell Plateau physiographic regions, 200 km north of Bowling Green, KY, and part of a contiguous karst landscape underlain by Mississippian-age carbonates. The Lost River and associated caves, sinkholes, sinking streams, and springs represent a world-class karst landscape and have been a subject of study since 1862 (Florea et al., 2018, and references therein). Approximately 4 km south of Orleans, the Lost River sinks during base flow and travels underground through mapped and
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
Flood Risk and Economic Exclusion
unmapped caves for a straight-line distance of 11 km to the True Rise of Lost River and likely Orangeville Rise, which are both second-magnitude springs (discharge greater than 10 ft3/s or 0.28 m3/s) and the second and third largest springs in Indiana, respectively (Blatchley, 1903). Orangeville Rise and Wesley Chapel Gulf (a karst window where the underground Lost River surfaces briefly) are listed as National Natural Landmarks. Waters rising from the True Rise of Lost River and Orangeville Rise are the headwaters of the Lost River downstream of the 37-km-long dry bed, the longest sinking stream in Indiana (Malott, 1952). The Lost River karst basin upstream of these springs is divided into two major sub-basins delineated by dye tracing: the Lost River sub-basin and the Orangeville Rise sub-basin (Bassett, 1974). The Lost River sub-basin primarily routes water from east of Wesley Chapel Gulf to the True Rise of the Lost River, a distance of about 14 km. The Orangeville Rise sub-basin primarily routes water from as far north as State Highway 60, some 10 km away, and from 4 km northeast of Orleans, a total distance of 13 km (Figure 2). State Highway 37 bisects both sub-basins, and the highway north of Orleans was expanded in 1976 to 4 lanes. In Orleans, an ephemeral surface-water channel, Flood Creek, collects water from a 41.8 km2 surface catchment composed mostly of agricultural fields. It is not represented as a flow line (the digitized extent of a perennial stream) on the National Hydrography Dataset (NHD). During normal flow conditions, Flood Creek is a sinking stream, and all flow is channeled underground in a sinkhole just north of Orleans that has been dye traced to Orangeville Rise (Bassett, 1974; Bayless et al., 1994). During flood events, the dry channel of Flood Creek is reactivated, and portions of Orleans flood with regular frequency (Malott, 1952; Bayless et al., 1994) (Figure 1). A flood-control structure was constructed 1 km north of town to impound runoff water and mitigate the impact of flash floods. While that impoundment alleviates some events, flooding remains a regular concern for residents of Orleans, including the segment of State Highway 37 through the center of town (Figure 1). Floods in the Lost River karst basin are not restricted to Orleans. Sinkholes across the Mitchell Plateau collect ponded stormwater and recharge the underlying conduit system. Concurrently, the dry bed of Lost River experiences annual flash floods, but often the water sinks into the caves through swallow holes in a matter of hours (Malott, 1931). Prolonged rainfall can cause severe flooding when the underground conduits fill and the dry bed reactivates in an upstream progression as the potentiometric surface rises. During these valley tides, the swallow holes in the dry bed reverse flow, water rises in sinkholes from below, and storm waters fill normally dry
valleys. While these flood hazards are known to residents, are well documented (Malott, 1952) (Figure 1), and even monitored (e.g., Bayless et al., 2014), they remain unmapped. MID-STATES CORRIDOR The Mid-States Corridor Project proposed two alternate routes for study that cut across the Crawford Uplands and the Mitchell Plateau. The draft EIS from this Phase I study identified north-south linkages between Owensboro, KY, and Indianapolis, IN, as a need for the corridor to connect industry and communities to Interstate Highways 64 and 69. The selected alternative, Route P, moved forward to the second phase of environmental review and engineering design (a Tier II EIS). This route would continue along U.S. Highway 231 to join Interstate 69 near Crane Naval Surface Warfare Center (Lochmueller Group, 2022). This route steered new highway construction west of the karst in the Mitchell Plateau. Northeast of Jasper, IN, Routes O and M were designed to connect to State Highway 37 south of Bloomington (Figure 2). While neither of these routes were selected for a Tier II EIS, the prospect of new highway construction through rural communities living on karst terrain raised interest and concern about the impacts that this and future highway infrastructure projects could have on people and the environment in a fragile landscape. Route O was of particular interest to this case study. The alignment of Route O bisected the Lost River karst basin (Figure 3). This route connected the town of Jasper through the Crawford Uplands to French Lick, IN, where sulfur-rich mineral springs were the focus of a historic spa and hotel (Blatchley, 1903). The hotel and resort were renovated and reopened in 2006 with a casino. North of French Lick, Route O entered the karst of the Mitchell Plateau, where it was located immediately adjacent to Orangeville Rise and the lower third section of the dry bed of the Lost River (Figure 3). Nine storm-water rises and two perennial rises of the Lost River and Orangeville Rise karst sub-basins were situated in the 2-mi-wide (3.2-km-wide) buffer for this proposed route (Figure 3). A karst window, the Ragdale Gulf, and the dye-traced conduit of Orangeville Rise were located directly beneath the center line of that proposed route (Figure 3). Major environmental concerns present along Route O included threats to threatened and endangered species, impacts to sustainable agriculture in the Amish communities and century farms in Lost River, and reductions to water quality from road runoff and hazardous spills. In these ways, the concerns are similar to those described in other controversial highway projects crossing karst
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
195
Burgess and Florea
Figure 3. Clyde Malott’s map of the dry bed of Lost River (Malott, 1952). The 0.61-km-wide Mid-States Corridor Alternative Route O is overlain as purple lines with an adjacent 1.61-km-wide buffer from the center line in light purple.
landscapes, such as the proposed Interstate 66 in southern Kentucky (Florea et al., 2002). The focus in this study, however, was on the ways in which flooding in the Lost River watershed could impact and be impacted by highway construction. This study was not a comprehensive investigation; it neither modeled the direct impacts of highway construction nor produced a predictive model on sections of proposed highway that are at increased risk. Rather, we used GIS inundation models to quantitatively evaluate the example of Orleans as a window into a broader issue of how current flood maps inaccurately portray karst landscapes and how this aspect of environmental justice is overlooked in those communities susceptible to flooding in karst. METHODS Flood inundation models were prepared to visualize and compare the threat of flooding to Orleans when modeled with and without consideration of karst flooding. The NHD Zone 05 digital elevation model (DEM) was clipped to the Lost River watershed as derived from Hydrologic Unit Code (HUC) 11 watershed boundaries with a 4 km buffer. This DEM included both the karst basin (upstream) and non-karst (downstream) portions of the watershed. Pits in the DEM (e.g., anomalous or missing elevations for pixels in the raster data) were filled to create a contiguous surface, and the TauDEM functions in ESRI ArcToolbox that use the D8 and D-inf algorithms were used to calculate the flow directions for each pixel for the watershed DEM. Initially, NHD flow lines were used to derive dangling points for a weighted 196
grid, which in turn was used to calculate the D8 contributing area. (The NHD line elements in the Lost River HUC 11 watershed are disconnected because of sinking streams.) A height above nearest drainage (HAND) flood inundation model was then constructed by calculating the vertical D-inf distance downstream (Figure 4) (Nobre et al., 2011). The U.S. Geological Survey water-level gauge on the Lost River near French Lick has a recorded peak stage between 7 and 8 m between 2010 and 2020 (U.S. Geological Survey, 2023) (Figure 2). For this reason, the HAND model was classified to 8 m to create a flood inundation model (Figure 5). Because Flood Creek is not included in the NHD, this process was repeated to construct a second model with additional dangling points added in Orleans at the upstream end of threshold-delineated paleo-valleys, creating a surface connection between Flood Creek and Lost River (Bayless et al., 2014). These inundation models were compared to the published Flood Insurance Rate Map (FIRM) data for the Lost River watershed (FEMA, 2014) (Figure 6). Finally, we acquired Landsat 7 satellite imagery of the Lost River karst basin from May 4, 2011 (Figure 7a). This followed the May 1–3, 2011, storm event, when the U.S. Geological Survey gauging station on the Lost River near French Lick measured a stage of 8.2 m, resulting in the flooding shown in Figure 2. Approximately 23 percent of the satellite image tile was missing data, 28 percent was obscured by cloud cover, and another 13 percent was muted by cloud shadow. The image area of the Orleans sub-catchment was mostly complete, and significant inundation was clearly visible
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
Flood Risk and Economic Exclusion
Figure 4. Height above nearest drainage (HAND) models for the Lost River watershed using only National Hydrography Dataset (NHD) flow lines (top) and another when the paleo-valley in the Orleans sub-catchment is connected to the Lost River (bottom).
as dark areas. A digitized flood area (Figure 7c) for the sub-catchment was compared to the HAND inundation model prepared at the same stage conditions. RESULTS The two HAND models (Figure 4) showed similar results for the majority of the Lost River watershed, with the exception of the Flood Creek sub-catchment. The addition of dangling points in Orleans allowed Flood Creek to be inundated (Figure 5), matching observations from both ground and satellite (Figure 7a). When this paleo-valley is considered active and connected to Lost River, a large proportion of Orleans can be inundated in the HAND models: from 5 percent to more than 60 percent when gauge height in the Lost River is 1 m and 8 m above base flow, respectively (Table 1). Furthermore, both models showed significant potential for flooding in the dry bed of the Lost River that is not considered in the FIRM 100-year-flood boundaries (Figure 6), including those areas that have been evaluated as a Special Flood Hazard Area (SFHA) by FEMA—defined as the area that will be inundated by the flood event having a 1 percent chance of being equaled or
Figure 5. Flood inundation models for the Lost River Watershed using only NHD flowlines (top) and when paleo valleys in the Orleans sub-catchment are connected to the Lost River (bottom). Gauge height is given in meters and is scaled based off the USGS gauging station on the Lost River near French Lick. The color ramp for each gauge height range matches the color shading of rows in Table 1.
exceeded in any given year. Alternative Route O of the Mid-States Corridor Project traversed portions of these potential areas of flooding. The HAND model inundated area (Figure 5) for the 8.2 m stage event of May 1–3, 2011 (27.6 km2 or 66 percent of the Orleans sub-catchment), over-predicted the actual flood area as digitized from Landsat 7 imagery on May 4, 2011 (Figure 7c). However, the imagery was acquired after the storm system and associated cloud cover cleared and likely underrepresents the peak inundation of Orleans. In short, the actual flood extent falls somewhere in between these two end members. The modeled inundation encompasses most of the satelliteobserved flood areas, which are expected to be constrained more to individual or connected sinkholes. DISCUSSION Floods are the most frequently occurring natural hazard and are exaggerated with climate change (NOAA,
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
197
Burgess and Florea
Figure 6. Map of the Lost River watershed (outlined in black) and the Orleans sub-catchment (outlined in red) with flood insurance risk map (FIRM) floodplain areas. Colored areas are 100 year floodplains. Red areas have been evaluated for Special Flood Hazard Area (SFHA). The remainder of the watershed is classified as a non-flood hazard.
2020). According to U.S. Congress, national losses due to flooding are rising (U.S. Code, 1973, Title 42, §4002). This observation, along with ongoing equity initiatives, has spurred regulatory reinvigoration of the NFIP under direction of FEMA’s Risk Rating 2.0 to update flood risk maps with SFHAs, where floodplain management regulations are strictly enforced, and flood insurance is mandatory. Karst communities can be resource limited by low- to medium-productivity soils, unreliable access to water, and a dynamic landscape that can be difficult for largescale infrastructure. While high-growth areas on karst exist in the United States (e.g., Florida and Texas), overall population growth on karst terrains is slow (Goldscheider et al., 2020). Rural areas in the lower Midwest, such as in Indiana and Kentucky, suffer from high rates of per capita income inequality (Ajilore and Willingham, 2020). The correlation between karst landscapes and rural populations in the lower Midwest contributes to this trend. Flooding and lack of access to the NFIP amplify this inequity. Previous work by the U.S. Geological Survey collected continuous stage monitoring data and some discharge measurements from areas around the Lost River watershed that were used to produce a Water Availability Tool for Environmental Resources (WATER) TOPMODEL (Bayless et al., 2014). The model combined closed-conduit and surfaceflow regimes in an attempt to capture the aquifer response of the upper Lost River watershed, but like many similar models, it has low correlation values with actual data. Although the tool was ostensibly developed to aid flood mapping in the Lost River watershed, no map was ever published. Similarly, according to the Indiana Department of Environmental Management’s 10-83 Watershed 198
Figure 7. (A) Landsat 7 satellite image of a portion of the Lost River watershed from May 2011. The sub-horizontal stripes are areas of missing image data. An example of an area of shadow from cloud cover is shown in the white circle. The white outline delineates the Orleans sub-catchment. (B) A portion of the Lost River watershed with the Orleans sub-catchment shown as a red outline, FIRM-designated areas for inclusion in the NFIP shown in blue, and SFHAs shown in yellow. (C) A portion of the Lost River karst basin with the Orleans sub-catchment shown as a red outline. The magenta polygon is the HAND-predicted flood area in the Orleans sub-catchment modeled at a stage of 8 m at the U.S. Geological Survey gauging station on the Lost River near French Lick (the white square in the lower left of all three tiles). The blue polygons are the outlines of flood-inundated areas of the image in the Orleans sub-catchment, which are visible as dark areas in the rest of the image.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
Flood Risk and Economic Exclusion Table 1. Provides total area inundated per flood gauge in the Lost River watershed calculated according to NHD flow lines and the Orleans catchment with paleo valley activation. The color shading of the table rows aligns with the color ramp in Figure 5 for each range in gauge height.
Flood Gauge (m)
Area Inundated
0-1 1-2 2-3 3-4 4-5 5-6 6-7 7-8
Lost River Lost River Orleans Orleans Watershed (km2) Watershed (%) Catchment (km2) Catchment (%) 18.60 1.96 2.06 4.93 40.61 4.28 4.46 10.68 62.19 6.56 6.91 16.53 83.01 8.75 9.85 23.56 107.92 11.38 13.18 31.51 133.13 14.04 16.35 39.10 157.08 16.56 20.11 48.09 210.50 22.20 27.63 66.07
Management Plan for Lost River (Korinek, 2013), the Indiana Department of Natural Resources may revisit flood boundaries using the U.S. Geological Survey tool or a different tool, but no revised flood map or update to include SFHAs has been published to date. The Lost River Watershed Flood Insurance Study (FIS) was most recently updated in 2014, but it did not consider Orleans for flood risk assessment (FEMA, 2014). Although damaging floods in Orleans occur on a regular basis, limited aerial photography of the floods is available to map the extent of inundation in detail during peak stage or to compare against simulations produced in this or other studies. The inundation models in this report (Figure 5) do not accurately represent flooding in the dry reaches of the Lost River because they do not account for the buffering capacity of the underground conduits—a key characteristic of karst landscapes. Since the utilized U.S. Geological Survey gauging station is downstream of the True Rise of the Lost River and Orangeville Rise (Figure 2), the inundation map certainly overestimates the area flooded per stage compared to expectations in French Lick. This inundation model also does not consider thousands of sinkholes and smaller sinking streams in the Lost River watershed that experience intermittent floods, and the method used may actually remove some of these features due to the pit-filling step used to derive the HAND models. The inundation area digitized from the Landsat 7 imagery illustrates this over-prediction and the lack of sinkhole-specific flooding (Figure 7). Still, the models clearly illustrate the following important points: (1) The inundation potential and magnitude of flooding in Orleans can be modeled when Flood Creek is connected to the NHD (Figure 5), and (2) flood extents are more expansive and more frequent than current FIRM models predict (Figure 6). Both statements are supported by satellite imagery (Figure 7). Future studies of this area must address the question of when and to what degree are paleo-valleys reactivated and
connected to the surface-water sections of the Lost River watershed. Policy Considerations Orleans, IN, requires a detailed flood risk assessment and revision of the flood maps that are used in planning and regulation. The need for and application of the NFIP and SFHAs for Orange County (including Orleans and French Lick/West Baden Springs) were referenced in the 2014 unified planning and zoning document, and they are parallel with the most recent FIS of the Lost River watershed (Figure 6) (FEMA, 2014) and the U.S. Geological Survey study of flood pulses in the Orangeville Rise sub-watershed (Bayless et al., 2014). Despite these efforts, Orleans remains excluded from formal consideration in the NFIP because Flood Creek is not in the NHD, and the modeling ignored the sinking stream segments and dry channels. In contrast, French Lick, with its renovated hotels and casino, is in a SFHA because it is downstream of the resurgences of the Lost River and has NHD flow lines (Figure 6). The results of our simple inundation model demonstrate a shortcoming of traditional flood-mapping techniques in karst landscapes and emphasize the need for a comprehensive reassessment of methods and related policy gaps. More importantly, we highlight a microcosm of a larger, national problem. The populations of French Lick/West Baden Springs and Orleans are about the same (FEMA, 2014). French Lick/West Baden Springs has access to the NFIP; Orleans does not. The difference in this access is neither the result of different ordinances nor attempts at rectifying the concern. In the end, the disparity seems to be due to differences in geology (presence or absence of karst) and economic forcing (private interests and tourism). Although it is difficult to quantify the precise extent of the issue, U.S. flood policy largely does not consider karst. At the federal level, FEMA flood-mapping guidance mentions karst once in a single paragraph in an amendment to calculations for general runoff (FEMA, 2019). This amendment does not account for runoff impoundment in sinkholes and sinking streams, resulting in critical underestimation of flood risk. Ongoing considerations in FEMA’s Risk Rating 2.0 for distance to waterbodies may enhance that concern, and the 2023 U.S. Supreme Court ruling on Sackett vs. U.S. EPA may make it even more difficult to include sinkholes and sinking streams as part of FEMA flood risk models. The responsibility may remain with individual states, where flood-management policies for karst are often nonexistent; however, some model ordinances have proposed runoff spill-over models (e.g., Currens, 2012). Many conversations regarding flood control in karst occur in counties or municipalities. The city of Bloomington in Monroe County, IN, 40 mi (64 km) north of
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
199
Burgess and Florea
Orleans in the same physiographic region, has a larger, more affluent tax base, with a robust consideration of karst in storm-water and flood ordinances. Those ordinances, with regional applicability, include guidance for runoff calculations for flood models and development restrictions in sinkholes, including setbacks and drainage easements (Bloomington Indiana, 2020). Karst research over the past half-century has spanned both Bloomington and Orleans, suggesting that the difference in policy is not for difference in data quantity. A more likely explanation is that Bloomington, home to Indiana University, has more access to scientific knowledge and more municipal resources than rural Orleans. Implementation of SFHAs and published FIRMs produces dual outcomes, and thus future implementation should include careful evaluation of both positive and negative repercussions. These designations can devalue property, reducing the overall tax base for a community proportional to the amount of floodplain identified (Frazier et al., 2020). However, rural, economically depressed communities in karst that experience regular flooding, such as Orleans, already have devalued property and have found mechanisms to share economic risk outside the regulatory sphere. SFHAs and FIRMs inclusive of karst could increase protection and resiliency, but only with accurate risk assessment. Underestimating flood risk, the present case, restricts access to the NFIP. On the other hand, overestimating flood risk may unnecessarily burden individuals by forcing NFIP participation. In Orleans, formal designations for flood hazards could change the outcome of proposed infrastructure development such as the Mid-States Corridor Project—decisions that could impact property value and plans for future community development and connectivity.
of this would be a great detriment to rural communities and homesteads that are excluded from NFIP participation. Perhaps a larger concern is the noticeable lack of detail on karst presented in the draft EIS by the Lochmueller Group (2022). In the buffer zone for Route O, 21 caves, 1 spring, and 22 sinkholes were noted, despite greater numbers of each that are known. Part of this discrepancy is inherent to the resources used for the assessment and illustrates the need for more comprehensive databases and emphasis on additional mapping. In the case of karst features, they derive from information provided by the Indiana Cave Survey, which only represents data collected and submitted by volunteers. For sinkholes, these data are based on GIS-determined centroids on a DEM that predates light detection and ranging (LiDAR) acquisition, thus only capturing sinkholes above a threshold depth. The draft EIS only cited one study on karst on the Mitchell Plateau by Hasenmueller et al. (2003) that focused on water quality at springs in Spring Mill State Park, which is not in the Lost River watershed, despite the wealth of publications with sitespecific details in the area of proposed impact (Florea et al., 2018, and references therein). In our reading of the draft EIS, there does not seem to be enough information on karst resources for decision makers to make an informed choice based on environmental risk assessments for those resources. While the decision in the draft EIS to select a route west of significant karst avoids many concerns outlined in this paper, the decision was not grounded in comprehensive data and could have gone the other way.
Impacts of Highway Development on Flooding in the Lost River Watershed
Communities in karst are often excluded from U.S. flood policy. This exclusion comes despite an increased exposure to flood risk, resultant damages to life and property, long-term consequences to agriculture and water quality, and susceptibility to water-borne illnesses. These impacts collectively decrease community and environmental resiliency. The hidden nature of caves makes comprehensive approaches to policy in karst genuinely difficult. Even so, what may seem like innocuous or even necessary assumptions about hydrology in karst can have significant repercussions for those who can and cannot participate in flood protection programs. The disconnect between scientists, locals, and policy makers has excluded populations across large geographic swathes from these economic spaces. Problems with property damage, agriculture, and water quality ensuing from floods compound economic and environmental justice issues. This burdens landowners, communities on karst, and the entire country.
Alternate Route O of the Mid-States Corridor Project avoided the Orleans catchment and significant interactions with current FIRM areas (Figure 2). However, the route alignment adjacent to the lower third of the dry bed of the Lost River (Figure 3) may exacerbate flooding in Orleans. For example, if clay, silt, fill, and trash from construction or runoff clog some of the nine stormwater rises in the dry bed in the buffer zone of alternative Route O, it is quite possible that the flood response of the Lost River drainage system will become more intensified (see also the failure of the Swift 2 Dam). Restricted storm-water rises could lead to higher pressure in conduits and backflooding of upstream swallow holes and sinkholes, perhaps as far upstream as Orleans (Figure 2). The situation may be even more dire if conduits collapse or are filled during construction, such as those between Ragdale Gulf and Orangeville Rise (Figure 3). All 200
CONCLUSIONS
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
Flood Risk and Economic Exclusion
An example of this environmental injustice is demonstrated in our flood inundation models of the Lost River watershed (Figure 5) that use a weighted grid with and without the sinking stream and contributing dry channels of Flood Creek surrounding Orleans, which are not included in the NHD. The results are comparative maps revealing areas with greater flood vulnerability. Without incorporating sinking streams and dry beds as potential conduits for surface-water flow, no risk of flood inundation appears in the Orleans catchment. This oversight means that policy makers do not consider flood hazards in Orleans. The inclusion of the ephemerally dry channels results in significant inundation despite the need for further model calibration using flood reports and storage characteristics of conduits; our models relied on a U.S. Geological Survey stream gauge downstream of the resurgence of the Lost River and thus overestimate flooded areas by inflating estimates of drainage in the upstream catchment. A satellite ortho-photo-derived inundation map during a May 2011 flood event illustrates the model’s over-prediction. Regardless, the models demonstrate the potential for the known and documented flooding in Orleans. According to U.S. Geological Survey stream gauge data from the Lost River, water levels rose with a frequency of 11 times between 1993 and 2014 to the 4–5 m stage during peak flow (U.S. Geological Survey, 2023). Without stream gauge data more proximal to the Orleans catchment, our inundation model predicts that 31.5 percent of that catchment could be inundated during floods at that same stage. Constructing a new highway corridor that would bisect this karst landscape could serve to amplify rather than mitigate the flood inundation potential, and thus exacerbate an environmental justice issue in this rural community. To that end, alternate Route O of the Mid-States Corridor Project would have been a strongly undesirable outcome. Despite a lack of comprehensive data in the draft EIS, the recommendation was to seek a route that largely avoids karst. The unmitigated vulnerability of flooding in karst and exacerbated inequity created by the NFIP are at odds with the vision set forth by the NFIP itself. To see a change in results, there must be a change in approach. A first suggestion would be to energize continued communication with policy makers to design karst-specific SFHAs. This regulatory approach risks being ineffective or creating even more inequity. A second suggestion would be to support broader mapping of karst resources, such as sinkholes, as proposed in the Sinkhole Mapping Act of 2021 considered by the U.S. Congress (U.S. Congress, 2021, H.R. 3681). More broadly, this dilemma emphasizes the wide collaborative gap between scientists and locals and the
knowledge gap between these groups and policy makers (Florea and Vacher, 2011). While karst scientists can technically explain how and why water moves underground, they do not always know where it will appear. Through lived experience and historical knowledge, locals have familiarity with where karst flooding occurs. Changing policy alone will likely not achieve desired outcomes without incorporating these perspectives. Rethinking community conversations about flooding with a focus on generating and validating community perspectives is critical to achieving more comprehensive flood hazard awareness, policy, and predictions. A collaborative approach to environmental justice in communities can imagine solutions beyond currently prescribed policy while building resiliency and partnerships. Instead of focusing energy toward the inertial barriers to systemic change in national policy, our communities may be best served by directing that energy inwards to address injustice. ACKNOWLEDGMENTS The authors value the input from three anonymous reviewers of an earlier draft of this manuscript and thank Environmental & Engineering Geoscience Co-Editor Eric Peterson for encouraging this contribution to this special issue on karst. The Indiana Geological and Water Survey provided a space for these conversations to develop into this paper, and the Center for Rural Engagement at Indiana University supported this work. The first author is grateful to Dr. Rebecca Lave and Dr. Taehee Hwang of the Indiana University Department of Geography for offering coursework that led to the set of ideas presented herein. REFERENCES AJILORE, O. AND WILLINGHAM, Z., 2020, The Path to Rural Resilience in America. Washington, D.C.: Center for American Progress. ALBÉRIC, P., 2004, River backflooding into a karst resurgence (Loiret, France): Journal of Hydrology, Vol. 286, No. 1–4, pp. 194–202. BAEDKE, S. J. AND KROTHE, N. C., 2001, Derivation of effective hydraulic parameters of a karst aquifer from discharge hydrograph analysis: Water Resources Research, Vol. 37, No. 1, pp. 13–19. BASSETT, J. L., 1974, Hydrology and Geochemistry of Karst Terrain, Upper Lost River Drainage Basin, Indiana: Master’s thesis, Indiana University, Bloomington, IN, 102 p. BAYLESS, E. R.; CINOTTO, P. J.; ULERY, R. L.; TAYLOR, C. J.; MCCOMBS, G. K.; KIM, M. H.; AND NELSON, H. L., 2014, Surface-Water and Karst Groundwater Interactions and Streamflow-Response Simulations of the Karst-Influenced Upper Lost River Watershed Orange County, Indiana: U.S. Geological Survey Scientific Investigations Report 2014-5028, 39 p. BAYLESS, E. R.; TAYLOR, C. J.; AND HOPKINS, M. S., 1994, Directions of Ground Water Flow and Locations of Ground-Water Divides in the Lost River Watershed near Orleans, Indiana: U.S. Geological Survey Water Resources Investigations Report 94-4195, 25 p. BLATCHLEY, W. S., 1903, The Mineral Waters of Indiana: Their Location, Origin and Character: W.B. Burford, contractor for state printing and binding, Indianapolis, IN.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
201
Burgess and Florea Bloomington Indiana, 2020, Unified Development Ordinance: City of Bloomington, IN. Electronic document, available at https:// bloomington.in.gov/sites/default/files/2020-04/Bloomington%20 UDO_04-18-2020_final.pdf BROWN, G. O., 2003, The history of the Darcy-Weisbach equation for pipe flow resistance. In Rogers, J. R. and Fredrich, A. J. (Editors), Environmental and Water Resources History: American Society of Civil Engineers, New York, pp. 34–43. CFR (Code of Federal Regulations), 2018, Title 44—Emergency Management and Assistance: Office of the Federal Register National Archives and Records Administration, Washington, D.C. CRAWFORD, N. C., 2005, Ground-water basin catchment delineation by dye tracing, water table mapping, cave mapping, and geophysical techniques: Bowling Green, Kentucky. In Yuhr, L. B.; Alexander, E. C., Jr.; and Beck, B. F. (Editors), Sinkholes and the Engineering and Environmental Impacts of Karst: American Society of Civil Engineers, New York, pp. 394–402. CURRENS, J. C., 2012, Model ordinance for development on karst terrain: Kentucky, USA: Carbonates and Evaporites, Vol. 27, No. 2, pp. 133–136. DOUGHERTY, P. H., 1983, Valley tides—Land-use response floods in a Kentucky karst region. In Dougherty, P. H. (Editor), Environmental Karst: Geospeleo Publications, Cincinnati, OH, pp. 3–15. FEMA (Federal Emergency Management Agency), 2014, Flood Insurance Study: Orange County, Indiana and Incorporated Areas: Flood Insurance Study Number 18117CV000A, FEMA, Washington, D.C. FEMA (Federal Emergency Management Agency), 2019, Guidance for Flood Risk Analysis and Mapping: General Hydrologic Considerations: FEMA, Washington, D.C. First Street Foundation, 2021, The 3rd National Risk Assessment: Infrastructure on the Brink: First Street Foundation, Brooklyn, NY, 163 p. FLOREA, L. J.; HASENMUELLER, N. R.; BRANAM, T. D.; FRUSHOUR, S. S.; AND POWELL, R. L., 2018, Karst geology and hydrogeology of the Mitchell Plateau of south-central Indiana. In Florea, L. J. (Editor), Ancient Oceans, Orogenic Uplifts, and Glacial Ice: Geologic Crossroads in America’s Heartland: Field Guide 51, Geological Society of America, Boulder, CO, pp. 95–112. FLOREA, L. J.; PAYLOR, R. L.; SIMPSON, L.; AND GULLEY, J., 2002, Karst GIS advances in Kentucky: Journal of Cave and Karst Studies, Vol. 64, No. 1, pp. 58–62. FLOREA, L. J. AND VACHER, H. L., 2011, Communication and forestructures at the geological intersection of caves and subsurface water flow—Hermeneutics and parochialism: Earth Sciences History, Vol. 30, No. 1, pp. 85–105. FRAZIER, T.; BOYDEN, E. E.; AND WOOD, E., 2020, Socioeconomic implications of national flood insurance policy reform and flood insurance rate map revisions: Natural Hazards, Vol. 103, pp. 329–346. GABROVŠEK, F.; PERIC, B.; AND KAUFMANN, G., 2018, Hydraulics of epiphreatic flow of a karst aquifer: Journal of Hydrology, Vol. 560, pp. 56–74. GOLDSCHEIDER, N.; CHEN, Z.; AULER, A. S.; BAKALOWICZ, M.; BRODA, S.; DREW, D.; HARTMANN, J.; JIANG, G.; MOOSDORF, N.; STEVANOVIC, Z.; AND VENI, G., 2020, Global distribution of carbonate rocks and karst water resources: Hydrogeology Journal, Vol. 28, No. 5, pp. 1661–1677. GULLEY, J.; MARTIN, J. B.; SCREATON, E. J.; AND MOORE, P. J., 2011, River reversals into karst springs: A model for cave enlargement in eogenetic karst aquifers: Geological Society of America Bulletin, Vol. 123, No. 3–4, pp. 457–467. HASENMUELLER, N. R.; COMER, J. B.; AND ZAMANI, D. D., 2003, Escherichia coli monitoring in the Spring Mill Lake watershed in southcentral Indiana. In Yuhr, L. B.; Alexander, E. C., Jr.; and Beck, B. F. (Editors), Sinkholes and the Engineering and Environmental
202
Impacts of Karst: American Society of Civil Engineers, New York, pp. 309–320. KEMMERLY, P., 1981, The need for recognition and implementation of a sinkhole-floodplain hazard designation in urban karst terrains: Environmental Geology, Vol. 3, pp. 281–292. https://doi.org/10. 1007/BF02473519 KIM, M., 2023, Two floods hit different states. One area got less money: E&E News. Electronic document, available at https://www.eenews .net/articles/two-floods-hit-different-states-one-area-got-less-money/ KIPPER, C., 2019, Influence of Spring Flow Reversals on Cave Dissolution in a Telogenetic Karst Aquifer, Mammoth Cave, KY: M.S. Thesis, Western Kentucky University, Bowling Green, KY. KORINEK, G., 2013, Watershed Management Plan for the Lost River Watershed: Lost River Watershed Partnership, Indiana Department of Environmental Management, Indianapolis, IN, 362 p. Lochmueller Group, 2022, Mid-States Corridor Draft Environmental Impact Statement: Lochmueller Group, Indianapolis, IN. Electronic document, available at https://midstatescorridor.com/deis/ MALOTT, C. A., 1931, Lost River at Wesley Chapel Gulf, Orange County, Indiana: Proceedings of the Indiana Academy of Science, Vol. 41, pp. 285–316. MALOTT, C. A., 1952, The swallow-holes of Lost River, Orange County, Indiana: Proceedings of the Indiana Academy of Science, Vol. 61, pp. 187–231. NOAA (National Oceanic and Atmospheric Administration), 2020, U.S. Resilience Toolkit: Inland Flooding: 25 March 2020, NOAA, Washington, D.C. NOBRE, A. D.; CUARTAS, L. A.; HODNETT, M. G.; RENNÓ, C. D.; SILVEIRA, A.; WATERLOO, M. J.; AND SALESKA, S., 2011, Height above the nearest drainage, a hydrological relevant new terrain model: Journal of Hydrology, Vol. 404, pp. 13–29. NWS (National Weather Service), 2010, United States Flood Loss Report—Water Year 2010: National Weather Service, Washington, D.C., p. 1 from the original on November 2, 2012. Electronic document, available at https://www.weather.gov/media/water/ WY10%20Flood%20Loss%20Summary.pdf PEKAŠ, Z.; JOLOVIĆ, B.; RADOJEVIĆ, D.; PAMBUKU, A.; STEVANOVIĆ, Z.; KUKURIĆ, N.; AND ZUBAC, Z., 2012, Unstable regime of Dinaric karst aquifers as a major concern for their sustainable utilization. In Proceedings of 39th International Association of Hydrogeologists Congress: International Association of Hydrogeologists, Niagara Falls, Canada, CD. PRYOR, S. C.; SCAVIA, D.; DOWNER, C.; GADEN, M.; IVERSON, L.; NORDSTROM, R.; PATZ, J.; AND ROBERTSON, G. P., 2014, Midwest. Climate change impacts in the United States: The third national climate assessment. In Melillo, J. M.; Richmond, T. C.; and Yohe, G. W. (Editors), National Climate Assessment Report: U.S. Global Change Research Program Washington, D.C., pp. 418–440. REGAN, P.; AND LANGNION, W. N., 2008, Failure of Swift 2 Forebay Dam (abstract). In 2008 United States Society on Dams Annual Meeting and Conference: U.S. Society on Dams, Portland, OR, p. 85. SIMPSON, L. C. AND FLOREA, L. J., 2009, The Cumberland Plateau of eastern Kentucky. In Palmer, A. N. and Palmer, M. V. (Editors), Caves and Karst of America: National Speleological Society, Huntsville, AL, pp. 70–79. U.S. Code, 1973, Chapter 50—National Flood Insurance. In Title 42— The Public Health and Welfare: U.S. Government Publishing Office, Washington, D.C., §4002, Pub. L, 93–234, §2, Dec. 31, 1973, 87 Stat. 975. U.S. Congress, 2021, H.R. 3681—The Sinkhole Mapping Act of 2021: Report No. 117–677, 117th Congress, 2nd Session. U.S. Geological Survey, 2023, National Water Information System Data: Electronic document, available at https://waterdata.usgs.gov/ monitoring-location/03373560
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 191–202
Hydrochemical Delineation of Spring Recharge in an Urbanized Karst Basin, Central Kentucky ALAN E. FRYAR* BENJAMIN J. CURRENS1 CRISTOPHER S. ALVAREZ VILLA2 Department of Earth and Environmental Sciences, University of Kentucky, 101 Slone Building, Lexington, KY 40506-0053
Key Terms: Hydrogeology, Karst, Kentucky, Non-Point Source Pollution, Stable Isotope, Stormwater
were effective in stormwater management during an unusually wet year.
ABSTRACT
BACKGROUND
Because of well-integrated surface and subsurface drainage in karst terrains, springs can exhibit relatively rapid hydraulic, chemical, and thermal responses to storms. In urbanized karst basins, impervious cover, stream channelization, and utility infrastructure can alter infiltration, provide alternate pathways for subsurface flow, and affect ambient water quality. We combined continuous logging of electrical conductivity (EC) and water temperature with analyses of stable isotopes (deuterium and oxygen-18) to differentiate focused and diffuse recharge in a karst basin in Lexington, Kentucky, during 2018. Logging occurred at the McConnell Springs Blue Hole and a sinkhole that drains to it; isotopes, specific conductance, and temperature were manually monitored at those sites and along two losing stream reaches. Water temperature at McConnell Springs and stable isotope abundances showed seasonal variability. The Blue Hole responded within hours to stormwater infiltration at the sinkhole (∼2.1 km upgradient), with recharge events marked by colder stormwater in winter and warmer stormwater in spring to early autumn. Stable isotopes indicated that sinkhole infiltration was minimally affected by evaporation during periods of ponding (up to 9 days). Spring discharge appeared to represent a mixture of focused and diffuse, partly evaporated recharge, consistent with a simple hydrologic model of rainfall, runoff, and infiltration in the basin. EC spikes at the spring during January–March were consistent with pulses of road salt or brine in runoff or snowmelt. Despite limited monitoring data, results suggest that restoration of the sinkhole and its inlet stream
Karst terrains are marked by integrated surface and subsurface drainage that links sinkholes or sinking streams to springs via conduit networks (White, 2002). Because of relatively rapid infiltration and flow along preferential pathways, spring discharge, temperature, and chemistry can exhibit pronounced responses to recharge events such as storms and snowmelt. Water, heat, and solutes can be exchanged between conduits and the rock matrix, and diffuse infiltration and flow through the matrix (e.g., via fractures) can also occur (Atkinson, 1977; Hess and White, 1988; and Luhmann et al., 2011). Urbanization can cause similar effects (“urban karst”; Bonneau et al., 2017). Impervious cover and stream channelization can accelerate runoff, thereby decreasing response time and increasing storm hydrograph peaks. In the subsurface, buried utility trenches can create preferential flow paths. Leakage from water mains augments recharge and infiltration to (or exfiltration from) storm and sanitary sewers can distort flow paths (Hibbs and Sharp, 2012; Passarello et al., 2012). Urban runoff and sanitary-sewer exfiltration can adversely impact water quality (Coakley et al., 2015; Robinson and Hasenmueller, 2017). Therefore, delineating sources and timing of recharge and contamination is complicated in urbanized karst terrains (Toran et al., 2009; Passarello et al., 2012; Robinson and Hasenmueller, 2017; Sharp, 2019; and Beal et al., 2020). The stable isotope (2 H and 18 O) composition of water can vary with sources and timing of precipitation and can be affected by partial evaporation, which differentially enriches the remaining liquid water in 18 O relative to 2 H (Darling et al., 2005). Consequently, 2 H and 18 O have become widely used in studies of recharge, including recharge to karst springs (e.g., Lakey and Krothe, 1996; Doctor et al., 2006; Ozyurt and Bayari, 2008; Jeelani et al., 2017; Robinson and Hasenmueller, 2017; Husic et al., 2019; and Zhang
1 Present address: Department of Plant and Soil Sciences, University of Kentucky, 105 Plant Science Building, Lexington, KY 40546 2 Present address: Cana Technology, 1250 45th Street, Suite 300, Emeryville, CA 94608 *Corresponding author email: alan.fryar@uky.edu
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
203
Fryar, Currens, and Alvarez Villa
et al., 2019). More recently, these isotopes have been employed to identify sources and travel times of water in urban basins (Soulsby et al., 2014; Ehleringer et al., 2016; Marx et al., 2021; and Stevenson et al., 2022). Sensors for real-time logging of electrical conductivity (EC) are also used to delineate the movement of storm pulses to karst springs. Such sensors enable measurement of changes in salinity associated with focused recharge, including both dilution by meteoric water and elevated solute concentrations in runoff (e.g., Hess and White, 1988; Liu et al., 2004; Toran et al., 2009; Mudarra et al., 2014; Reisch and Toran, 2014; Jeelani et al., 2017; Robinson and Hasenmueller, 2017; and Luo et al., 2018). In this study, we combined analyses of stable water isotopes and continuous EC logging to differentiate focused and diffuse recharge in an urbanized karst basin. We are aware of only one similar prior study (Robinson and Hasenmueller, 2017). Our work is distinguished by use of EC logging at both a sinkhole and a spring to constrain the timing of recharge pulses and the duration of sinkhole ponding following storms as well as by comparison of stable isotopic signatures at multiple sinks and the spring. These data, together with water temperature logging at the spring, meteorological data, and simplified rainfall-runoff modeling, enabled us to address the following question: how effectively does the modified karst drainage network function in stormwater management? STUDY AREA SETTING The study area is the 12.1-km2 McConnell Springs basin, located in Lexington, Kentucky (Figure 1). At least four focused recharge sites, including sinkholes and a swallet (streambed sink), have been identified by qualitative dye tracing (Currens et al., 2002). Most of the basin lies within the Wolf Run watershed (Figure S1, Appendix A; https://www.aegweb.org/e-egsupplements), but the McConnell Springs karst window lies within (yet is hydraulically isolated from) the adjoining Town Branch watershed. Groundwater discharges initially at the Blue Hole spring (Figure 2A), flows overland ∼40 m, sinks and flows underground ∼60 m to an artesian spring, flows overland ∼140 m and sinks again, and then finally discharges ∼900 m WNW at Preston’s Cave Spring, which is the headwaters of an unnamed perennial tributary to Wolf Run (Barton et al., 2012; Price, 2022). In the Inner Bluegrass region, preferential dissolution along joints and bedding planes in Ordovician limestone has created fluviokarst features (Phillips et al., 2004). Groundwater basins have formed where conduit development has pirated surface stream flow (Barton et al., 2012). The McConnell Springs basin is
204
underlain by the Tanglewood Limestone and the Brannon Member of the Lexington Limestone in upland areas and by the lower Lexington Limestone elsewhere (Kentucky Geological Survey, 2022). Soils formed as residuum from the weathering of shaly and argillaceous facies. The McConnell Springs basin is mantled largely by Bluegrass-Maury silt loam (2–6 percent slope, well drained) (Soil Survey Staff, 2023b). The landscape is gently rolling; total relief within the Wolf Run watershed is ∼60 m. The Inner Bluegrass region falls within KöppenGeiger climate classification Cfa (humid subtropical; Brugger, 2017), with no pronounced wet or dry seasons. In this area of the midcontinental United States, precipitation originates from the Pacific Ocean, the Gulf of Mexico, and continental and (to a lesser extent) Arctic sources, with proportions of precipitation from each source varying seasonally (Bedaso and Wu, 2020). For the U.S. National Weather Service station at Lexington Blue Grass Airport (WBAN 93820), the 20year average annual rainfall preceding the study (1998– 2017) was 1,236 mm, with May and July typically the wettest months. Daily air temperature ranged from an average low of −3.7°C in January to an average high of 29.9°C in July (Midwest Regional Climate Center [MRCC], 2022). Within the McConnell Springs basin, the largest land use/land cover categories are developed/low intensity (29 percent) and developed/open space (28 percent) (Table S1, Appendix B; https:// www.aegweb.org/e-eg-supplements), based on the 2011 National Land Cover Database (NLCD; Multi-Resolution Land Characteristics Consortium [MRLCC], 2022). Single-family housing is dominant, but there are also apartments, light commercial developments, three hospitals, parks (including McConnell Springs and the Picadome golf course), and schools (including part of the University of Kentucky [UK] campus). Natural karst drainage networks are incorporated into stormwater management in the basin. Notably, stormwater infiltrates via a sinkhole behind the Campbell House Inn on the Picadome golf course, ∼2.1 km SSE of the McConnell Springs Blue Hole (Figure 2B and Figure S2, Appendix A; https://www.aegweb.org/e-eg-supplements). This sinkhole drains the ephemeral Big Elm tributary, whose >2.0-km2 watershed includes ∼1.3 km2 of impervious cover; 38 mm of rainfall in 24 hours can cause flash flooding and overflow across the golf course to Vaughn’s Branch, a tributary to Wolf Run. Because of debris accumulation and scouring of the channel to bedrock, the sinkhole was excavated, and the tributary reach immediately upstream (165 m length) was restored in 2016 (Mehlhorn, 2016).
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
Hydrochemical Delineation of Spring Recharge in an Urbanized Karst Basin, Central Kentucky
Figure 1. (A) Inner Bluegrass region of Kentucky; red rectangle indicates area shown in (B) (modified from Barton et al., 2012); (B) McConnell Springs basin map with dye-trace vectors, sampling sites, USGS gauge, and impervious cover.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
205
Fryar, Currens, and Alvarez Villa
Figure 2. Photos of (A) the Blue Hole at McConnell Springs (February 14, 2020) and (B) the Campbell House sinkhole (center marked by red arrow) after 32 mm of rainfall in 24 hours (February 16, 2018).
206
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
Hydrochemical Delineation of Spring Recharge in an Urbanized Karst Basin, Central Kentucky
Infiltration capacity must be exceeded at two swallets along this reach before flow into the sinkhole can occur. METHODS We collected grab samples in 10-ml glass vials for water-isotope analyses from February 2 through December 12, 2018 (typically weekly) at four sites: the McConnell Springs Blue Hole, the Campbell House sinkhole, and along losing reaches of Vaughn’s Branch and Wolf Run (Figure 1B). The Campbell House sinkhole and the Wolf Run reach have been dye-traced to McConnell Springs (Currens et al., 2002; Garrison, 2019). Samples were collected at intervals ranging from 3 to 21 days, depending in part on when water was present at sites other than McConnell Springs. Samples were refrigerated at 5°C and filtered through 0.45-µm filters prior to analysis on a Los Gatos T-LWIA-45-EP liquid water isotope analyzer at the Kentucky Stable Isotope Geochemistry Laboratory (KSIGL). Analytical protocols are detailed in Avery et al. (2022). Values are reported in δ notation relative to Vienna Standard Mean Ocean Water–Standard Light Antarctic Precipitation, with long-term standard deviations of 0.2‰ for δ2 H and 0.1‰ for δ18 O (KSIGL, 2022). We measured specific conductance (SC; EC corrected to 25°C) and water temperature using a YSI Professional Plus meter (Xylem, Yellow Springs, OH) during most instances when isotope samples were collected from February 9 through December 12, 2018. Statistical analyses of manually measured parameters and isotope data were performed using R software (R Core Team, 2019) except for linear regressions, which were performed using Microsoft Excel. At the Campbell House sinkhole, EC was logged every 15 minutes from May 2 to October 12, 2018. At the McConnell Springs Blue Hole, EC and water temperature were recorded every 15 minutes from January 19 to March 1, March 9 to April 20, and May 2 to October 12, 2018. Both sites were monitored using Aqua TROLL 200 loggers (In-Situ, Fort Collins, CO), which have an EC range of 0–100,000 μs/cm with an accuracy of ±0.5 percent + 1 μs/cm and a calibrated temperature range of 0–50°C. At the first sink beyond the Blue Hole at McConnell Springs, water temperature was also recorded using a Hobo U22-001 logger (Onset Computer, Bourne, MA) every 15 minutes from January 28 to December 27, 2016. Ancillary data for calendar year 2018 include daily rainfall and hourly air temperature from Blue Grass Airport (MRCC, 2022) and 15-minute discharge data from U.S. Geological Survey (USGS) gauge 03289193 on Wolf Run (∼1.0 km downstream from the con-
fluence with the Preston’s Cave Spring tributary and ∼1.0 km upstream from the confluence with Town Branch) (USGS, 2022). The annual discharge hydrograph was separated into runoff and baseflow components using the Web-based Hydrograph Analysis Tool (WHAT, 2022) with the specified value of the maximum baseflow index for perennial streams with porous aquifers (0.80) (Lim et al., 2005). We ran the Model My Watershed Site Storm Model (Stroud Water Research Center, 2022) to calculate partitioning among evapotranspiration, infiltration, and runoff for a 25-mm rainfall in a 24-hour period over the Wolf Run watershed upstream of the confluence with the Preston’s Cave Spring tributary (Figure S1). This model uses the Source Loading and Management Model algorithm (SLAMM; PV and Associates, 2019) with the 2011 NLCD (MRLCC, 2022) and the gridded Soil Survey Geographic (gSSURGO) database (Soil Survey Staff, 2023a). RESULTS Meteorological and Stream Flow Data With 1,828 mm of precipitation (47.9 percent higher than the preceding 20-year average), 2018 was the wettest year since continuous weather records began in Lexington in 1888 (MRCC, 2022). February and September were the wettest months in 2018 and were wetter than those same months in 1998–2017, as was October. Maximum single-day rainfall during 2018 was 87 mm on October 4. Air temperature in 2018 ranged from an average low of −5.8°C in January to an average high of 30.7°C in July. April was cooler and May was warmer than in the preceding 20 years. Discharge at the Wolf Run gauge ranged from 111 liters per second (LPS) on July 14 to 10,340 LPS on September 9, 2018 (Figure S3, Appendix A; https://www.aegweb.org/e-eg-supplements), with a median of 487 LPS and a mean of 834 LPS. Relative to 1997–2017, mean discharge for February 2018 was higher than for any other February, and mean discharge for September was the second highest for that month. Mean water-year discharge (October 1, 2017– September 30, 2018) was 722 LPS, the highest since record keeping began in October 1997. According to WHAT calculations, baseflow constituted 52.5 percent of total discharge for calendar year 2018. Manual Field Parameters Median SC values were similar for Wolf Run (503 µs/cm, N = 32), the McConnell Springs Blue Hole (532 µs/cm, N = 40), and the Campbell House sinkhole (533 µs/cm, N = 23) but were notably
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
207
Fryar, Currens, and Alvarez Villa Table 1. δ2 H versus δ18 O regressions for monitoring sites in this study, Husic et al. (2019), and Kendall and Coplen (2001); N = number of measurements. Location
Date Range
Regression Equation
r2
N
Reference
Campbell House McConnell Springs Blue Hole Vaughn’s Branch Including outlier Without outlier Wolf Run Cane Run Royal Spring Royal Spring conduit Kentucky streams (6)
2/2/18–11/30/18 2/2/18–12/12/18
δ2 H = 7.27δ18 O + 6.86 δ2 H = 5.58δ18 O − 4.37
0.926 0.860
23 41
This study This study
2/2/18–12/12/18 2/2/18–12/12/18 2/2/18–12/7/18 10/8/17–10/11/17 10/8/17–10/16/17 10/7/17–10/16/17 1984–1987
δ2 H = 7.01δ18 O + 3.68 δ2 H = 6.18δ18 O − 0.910 δ2 H = 6.05δ18 O − 2.57 δ2 H = 7.55δ18 O + 0.547 δ2 H = 5.45δ18 O − 10.6 δ2 H = 6.62δ18 O − 0.934 δ2 H = 6.4δ18 O + 1.2
0.952 0.927 0.849 0.975 0.914 0.949 0.96
40 39 33 6 9 10 66
This study This study This study Husic et al. (2019) Husic et al. (2019) Husic et al. (2019) Kendall and Coplen (2001)
higher (738 µs/cm, N = 42) for Vaughn’s Branch (Table S2, Appendix B; https://www.aegweb.org/eeg-supplements). Except for Wolf Run, the highest SC values at each site occurred on March 23, 2018. Vaughn’s Branch had the highest SC value for 35 of 42 monitoring dates (by as much as 577 µs/cm) as well as the highest overall value and the broadest range of values (149–1,344 µs/cm). Minimum water temperature values were 7.2–8.5°C, and maximum values were 24.2–27.2°C for surface-water monitoring sites (Table S2). The Blue Hole had the narrowest water temperature range (11.9–22.3°C; median 16.2°C). Values of SC were normally distributed, but water temperature values were not; one-way analysis of variance and Tukey honest significant difference tests on mean values for SC and the non-nonparametric KruskalWallis test on mean values for temperature were performed for dates when parameters were measured at all sites (N = 23). Results indicated a significant difference (α = 0.05) between SC at Vaughn’s Branch and other sites but no significant difference in water temperature among sites. Isotopes Median stable isotope ratios were lowest for the Campbell House (δ2 H −41.2‰, δ18 O −6.6‰) and highest for Vaughn’s Branch (δ2 H −36.5‰, δ18 O −5.7‰). Ranges of δ2 H and δ18 O were narrowest for the McConnell Springs Blue Hole (−53.4 to −27.6‰ and −8.0 to −4.3‰, respectively) and broadest for Vaughn’s Branch (δ2 H −96.2 to −22.9‰ and δ18 O −13.2 to −3.7‰), which had the lowest minimum values overall (Table S2). Minimum values occurred in March at each site except the Blue Hole, for which the δ2 H minimum occurred on May 31 (Figure 3). Maximum values at each site occurred between August and October, with overall maximum values occurring at the Campbell House (δ2 H −12.6‰, δ18 O −1.8‰) on September 21. We did not find significant differ-
208
ences in mean δ2 H or δ18 O among the four sampling sites (based on the Kruskal-Wallis test) or in median δ2 H or δ18 O for pairs of sampling sites (based on the Mann-Whitney test). We also did not find significant differences in mean δ2 H or δ18 O based on the KruskalWallis test when samples were separated into baseflow and stormflow populations following WHAT analyses. Overall, stable isotope values tended to fall below the global meteoric water line (GMWL, δ2 H = 8δ18 O + 10‰; Craig, 1961) on a plot of δ2 H versus δ18 O (Figure 4). We do not have information on the isotopic composition of precipitation in Lexington, but local meteoric water lines for sites within 240 km (in southwest Ohio, central Indiana, and east Tennessee) had slopes slightly less than or equal to the GMWL (7.61–8.0; Coplen and Huang, 2000; Tian et al., 2018; Bedaso and Wu, 2020; and Smith et al., 2021). Linear regressions of δ2 H versus δ18 O values for each site had slopes ranging from 5.58 at the McConnell Springs Blue Hole to 7.27 at the Campbell House (Table 1). The upper 95 percent limit of the slope for the Blue Hole was greater than the slope for Wolf Run but less than the lower 95 percent limit of the slopes for the Campbell House and Vaughn’s Branch. However, because the minimum δ2 H and δ18 O values for Vaughn’s Branch were extreme outliers (National Institute of Standards and Technology, 2022), we recalculated the linear regression excluding the outlying data point and obtained a slope of 6.18 (Table 1), which was less than the upper 95 percent limit of the slope for the Blue Hole. Logged Electrical Conductivity and Water Temperature Logged EC values at the Campbell House ranged from 0 µs/cm (when the sensor was dry) to 818 µs/cm (on October 1, 2018) (Figure 5). Based on EC values ࣙ90 µs/cm, we identified 35 sinkhole ponding events between May 5 and October 11, with dura-
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
Hydrochemical Delineation of Spring Recharge in an Urbanized Karst Basin, Central Kentucky
Figure 3. (A) δ2 H versus time and (B) δ18 O versus time at monitoring sites (C = Campbell House, M = McConnell Springs Blue Hole, V = Vaughn’s Branch, W = Wolf Run). Stormflow and baseflow samples were classified using hydrograph separation at the Wolf Run stream gauge (USGS, 2022; WHAT, 2022).
tions from 2.25 to 212.75 hours (median 21.5 hours) (Figure 6). Logged EC at the McConnell Springs Blue Hole ranged from 176 µs/cm on May 5 to 2,937 µs/cm on February 7, 2018 (Figure 7). Apart from a storm
event on January 27–28, EC at the Blue Hole was >1,000 µs/cm from the start of logging on January 19 until February 1. Baseline EC gradually declined from ∼750 to 650 µs/cm between May 1 and October 12,
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
209
Fryar, Currens, and Alvarez Villa
Figure 4. δ2 H versus δ18 O relative to global meteoric water line (GMWL) at monitoring sites (C = Campbell House, M = McConnell Springs Blue Hole, V = Vaughn’s Branch, W = Wolf Run). Note that the depleted outlier for Vaughn’s Branch (δ2 H −96.2‰, δ18 O −13.2‰) does not appear at the scale of this graph.
when logging ended. Logged water temperature at the Blue Hole ranged from 6.17°C on March 24 to 23.75°C on June 12, 2018 (Figure 7). Baseline water temperature gradually increased from ∼12°C to 17°C between March and October. We identified 52 events at the McConnell Springs Blue Hole with an EC change ࣙ100 µs/cm from baseline values. Of these events, 49 occurred on days with measurable runoff identified by WHAT at the Wolf Run gauge, and 44 had a water temperature change ࣙ1.0°C. The EC minimum (or maximum) preceded the temperature minimum (or maximum) in 35 of 52 events, with the lag ranging from 0 to 17.5 hours overall. In January–March 2018, EC rose in 12 of 18 events (by a maximum of 1843 µs/cm), and temperature fell in 17 of 18 events (by a maximum of 7.08°C). During April–October, EC fell in all 34 events (by a maximum of 575 µs/cm), and temperature rose in 32 of 34 events (by a maximum of 7.15°C). Based on the onset of EC rise at the Campbell House and EC fall at the McConnell Springs Blue Hole (Figure 5), we inferred 35 recharge events (with the caveat that not all ponding events at the Campbell House resulted in an EC response at the Blue Hole). The elapsed time in EC responses between the sites ranged from 1.64 to 16.14 hours (median 6.14 hours) (Figure 6). Separately, during episodes of ponding at the Campbell House, we inferred 20 recharge events based on sequential EC minima at the Campbell House and the Blue Hole (11 ponding events had multiple minima). The elapsed time in EC responses for these events ranged from 2.39 to 23.14 hours (median 4.39 hours) (Figure 6).
210
DISCUSSION During the study period, storm-event responses were superposed on seasonal variability in water temperature and EC. Recharge events tended to be marked at the McConnell Springs Blue Hole by colder stormwater in January–March and warmer stormwater in April–October. These events were associated with daily rainfall totals ࣙ4 mm except for July 3, 5, and 7 (dates for which WHAT did not calculate any runoff at the Wolf Run gauge) and October 11. We observed broadly sinusoidal fluctuations of water temperature and stable isotope abundances (particularly δ18 O) during the year, with minima in late winter/spring and maxima in late summer/autumn. Because logging during 2018 occurred only until midOctober, we have included the logged 2016 data and manual 2018 temperature measurements from McConnell Springs to illustrate annual-scale variability in water temperature (Figure S4, Appendix A; https://www.aegweb.org/e-eg-supplements). Because of heat exchange with the matrix during groundwater flow (Luhmann et al. 2011, 2012), at the Blue Hole, stormwater temperature minima or maxima tended to lag EC minima or maxima, baseline minimum and maximum water temperatures lagged air temperatures by ∼1–2 months, and the water-temperature range was narrower than at surface-water monitoring sites. Electrical conductivity logging indicated relatively rapid infiltration at the Campbell House sink and flow to the McConnell Springs Blue Hole. Ponding lasted up to 9 days, but the relatively high slope
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
Hydrochemical Delineation of Spring Recharge in an Urbanized Karst Basin, Central Kentucky
Figure 5. (A) Logged EC at the Campbell House sinkhole and McConnell Springs Blue Hole with Lexington daily rainfall (May 2–October 12, 2018); (B) inset showing detail of EC variations at the Campbell House sinkhole and McConnell Springs Blue Hole (June 9–July 19, 2018).
of the δ2 H versus δ18 O regression at the Campbell House indicates minimal evaporative enrichment, in part because longer-duration ponding events were sustained by multiple storms. With sufficient time, EC values at the Campbell House approached baseline values at the Blue Hole, perhaps because of waterrock interaction during slow drainage (Desmarais and Rojstaczer, 2002). EC responses indicated flow to the
Blue Hole over periods typically ࣘ10 hours, consistent with dye traces showing first-arrival times of 3–6 hours (Garrison, 2019). Response times <3 hours may reflect local run-on to the Blue Hole or infiltration at sinks closer than the Campbell House to the Blue Hole, whereas response times >6 hours may reflect contributions from sinks that are farther away or less well connected to the conduit network.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
211
Fryar, Currens, and Alvarez Villa
Figure 6. Ponding durations at the Campbell House sinkhole and response times at the McConnell Springs Blue Hole to EC rises and minima at the Campbell House sinkhole.
Discharge at McConnell Springs represents a combination of quick flow through conduits and diffuse flow, which is consistent with model results for the Wolf Run watershed. For a hypothetical 25-mm rain event, 15.2 mm infiltrates, 6.8 mm runs off (some of which is diverted into the subsurface through sinks), and 3.1 mm is subject to evapotranspiration. Hydrologic soil group B in gSSURGO (moderate infiltration) occupies 67% of the modeled area. The relatively low slope of the δ2 H versus δ18 O regression at the McConnell Springs Blue Hole indicates partly evaporated recharge indicative of diffuse infiltration. By comparison, Husic et al. (2019) collected data during an October 2017 tropical cyclone in the nearby Royal Spring basin, which discharges in Georgetown, Kentucky (Figure 1A). The δ2 H versus δ18 O slope was lower for Royal Spring (5.45) than for a karst conduit feeding it (6.61) or the ephemeral reach of Cane Run (7.55) that recharges the conduit (Table 1). Likewise, for a karst basin in southwest China, Zhang et al. (2020) obtained a lower δ2 H versus δ18 O slope for groundwater (3.37) than surface water (5.28). McConnell Springs is susceptible to non-point source pollutants in runoff (Brion, 2011; Dixon et al., 2015; and Price, 2022). January–March EC spikes were concurrent with road salt or brine application, which occurred within 2 days prior to each EC spike except January 23 (and had occurred January 12– 19) (Dixon, 2022). Major streets and roads in Lexington are treated before and during snow and ice storms with salt, brine, or a 10:10:80 mixture of brine, BeetHeet (organic-based liquid deicer), and wa-
212
ter (Allen, 2022; Lexington-Fayette Urban County Government [LFUCG], 2022). The decrease in baseline EC during summer and fall is consistent with progressive flushing of residual salt from soil (Robinson and Hasenmueller, 2017). Excluding the exceptionally depleted outlier, the δ2 H versus δ18 O slope for the Vaughn’s Branch monitoring site (6.18) was similar to the Wolf Run site (6.05). Given that flow was perennial at the Vaughn’s Branch site (in contrast to the Wolf Run site), we infer that flow may have been sustained locally by sanitary-sewer or water-main leaks. Sanitary-sewer leakage has been detected elsewhere in the Wolf Run basin (Brion, 2011; Coakley et al., 2015), and dry-weather flow to a storm sewer on the UK campus was attributed to leakage from building drains and chilled-water lines (Lewis, 2018). We are uncertain whether the generally higher SC values along Vaughn’s Branch reflected elevated solute concentrations associated with leakage or greater salt/brine application in that part of the basin. Lexington’s water supply is taken from pumping stations on the Kentucky River and a local reservoir (Kentucky American Water Company, 2022). The revised δ2 H versus δ18 O slope for Vaughn’s Branch, which is slightly less than the composite value for six Kentucky streams (6.4; Table 1; Kendall and Coplen, 2001), may be consistent with leakage of treated river water. The outlier data point for Vaughn’s Branch, which coincided with relatively low SC, probably reflects rainfall and runoff; 6.4 mm fell on the sampling date (March 30), and 35.2 mm fell during the preceding 2 days.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
Hydrochemical Delineation of Spring Recharge in an Urbanized Karst Basin, Central Kentucky
Figure 7. Logged EC and water temperature at the McConnell Springs Blue Hole: (A) January 19–April 20, 2018; (B) May 1–October 12, 2018. Note change in y-axis scales between figures.
CONCLUSIONS Using logging of EC and water temperature, we have shown that an urbanized karst spring responds within hours to infiltration of stormwater at a sinkhole ∼2.1 km upgradient. From analyses of δ2 H and δ18 O, we infer that sinkhole infiltration is minimally affected by evaporation over periods of hours to days and that spring discharge represents a mixture of focused and diffuse (partly evaporated) recharge, consistent with a simple hydrologic model of rainfall, runoff, and infil-
tration in the basin. EC spikes at the spring during January–March are consistent with pulses of road salt or brine in runoff or snowmelt, whereas frequently elevated SC values along a losing stream upgradient of the spring may reflect sanitary-sewer or water-main leaks. Although we lack pre-restoration data, our results suggest that restoration of the sinkhole and its inlet stream were effective in stormwater management during an unusually wet year. Monitoring at a higher temporal resolution over a longer period, with measurements of δ2 H and δ18 O of rainfall, stage and flow
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
213
Fryar, Currens, and Alvarez Villa
data at monitoring sites, and synchronous dye tracing from multiple sinks, would facilitate determining the amounts of recharge from different sources. ACKNOWLEDGMENTS This project was supported in part by a National Science Foundation Graduate Research Fellowship to B. J. Currens. We appreciate the assistance of LFUCG Parks and Recreation in providing access, LFUCG Streets and Roads in providing data, Jordon Munizzi with isotope analyses, Ryan Dapkus with statistical analyses, and three anonymous reviewers in providing suggestions for revisions. Unpublished data are available from A. E. Fryar on request. REFERENCES Allen, R., 2022, personal communication, Department of Streets and Roads, Lexington-Fayette Urban County Government, Lexington, KY. Atkinson, T. C., 1977, Diffuse flow and conduit flow in limestone terrain in the Mendip Hills, Somerset (Great Britain): Journal Hydrology, Vol. 35, pp. 93–110. Avery, E.; Samonina, O.; Kryshtop, L.; Vyshenska, I.; Fryar, A. E.; and Erhardt, A. M., 2022, Use of isotopes in examining precipitation patterns in north-central Ukraine: Isotopes Environmental Health Studies, Vol. 58, pp. 380–401. Barton, A. M.; Black Eagle, C. W.; and Fryar, A. E., 2012, Bourbon and springs in the Inner Bluegrass region of Kentucky. In Sandy, M. R. and Goldman, D. (Editors), On and around the Cincinnati Arch and Niagara Escarpment: Geological Field Trips in Ohio and Kentucky for the GSA North-Central Section Meeting, Dayton, Ohio, 2012: Field Guide 27, Geological Society of America, Boulder, CO, pp. 19–31. Beal, L.; Senison, J.; Banner, J.; Musgrove, M. L.; Yazbek, L.; Bendik, N.; Herrington, C.; and Reyes, D., 2020, Stream and spring water evolution in a rapidly urbanizing watershed, Austin, TX: Water Resources Research, Vol. 56, e2019WR025623. https://doi.org/10.1029/2019WR025623 Bedaso, Z., and Wu, S.-Y., 2020, Daily precipitation isotope variation in Midwestern United States: Implication for hydroclimate and moisture source: Science Total Environment, Vol. 713, No. 136631. Bonneau, J.; Fletcher, T. D.; Costelloe, J. F.; and Burns, M. J., 2017, Stormwater infiltration and the “urban karst”—A review: Journal Hydrology, Vol. 552, pp. 141–150. Brion, G. M., 2011, A Plan for Identifying Hot-Spots and Affirming Remediation Impacts on Surface Water Quality: Phase I: Electronic document, available at http://www.wolfrunwater. org/2010-Wolf-Run-DNA-Pathogen-Brion2011.pdf Brugger, K., 2017, World Map of the Köppen-Geiger Climate Classification Updated Map for the United States of America: Electronic document, available at http://koeppen-geiger.vuwien.ac.at/usa.htm Coakley, T. L.; Brion, G. M.; and Fryar, A. E., 2015, Prevalence of and relationship between two human-associated DNA biomarkers for Bacteroidales in an urban watershed: Journal Environmental Quality, Vol. 44, pp. 1694–1698. Coplen, T. B. and Huang, W., 2000, Stable Hydrogen and Oxygen Isotope Ratios for Selected Sites of the National Oceanic and Atmospheric Administration’s Atmospheric Integrated Re-
214
search Monitoring Network (AIRMoN): U.S. Geological Survey Open-File Report 00-270. Craig, H., 1961, Isotopic variations in meteoric waters: Science, Vol. 133, pp. 1702–1703. Currens, J. C.; Paylor, R. L.; and Ray, J. A., 2002, Mapped Karst Ground-Water Basins in the Lexington 30 × 60 Minute Quadrangle: Series 12, Map and Chart 10, Kentucky Geological Survey, Lexington, KY, scale 1:100,000, 1 sheet. Darling, W. G.; Bath, A. H.; Gibson, J. J.; and Rozanski, K., 2005. Isotopes in water. In Leng, M. J. (Editor), Isotopes in Palaeoenvironmental Research: Springer, Dordrecht, The Netherlands, pp. 1–66. Desmarais, K. and Rojstaczer, S., 2002, Inferring source waters from measurements of carbonate spring response to storms: Journal Hydrology, Vol. 260, pp. 118–134. Dixon, A., 2022, personal communication, Department of Streets and Roads, Lexington-Fayette Urban County Government, Lexington, KY. Dixon, N.; Sanchez, C.; Fryar, A. E.; and Bandy, A. M., 2015, Water quality in the Wolf Run basin, Lexington, Kentucky: Geological Society of America Abstracts Programs, Vol. 47, No. 7, p. 526. Doctor, D. H.; Alexander, E. C., Jr.; Petrič, M.; Kogovšek, J.; Urbanc, J.; Lojen, S.; and Stichler, W., 2006, Quantification of karst aquifer discharge components during storm events through end-member mixing analysis using natural chemistry and stable isotopes as tracers: Hydrogeology Journal, Vol. 14, pp. 1171–1191. Ehleringer, J. R.; Barnette, J. E.; Jameel, Y.; Tipple, B. J.; and Bowen, G. J., 2016, Urban water—A new frontier in isotope hydrology: Isotopes Environmental Health Studies, Vol. 52, pp. 477–486. Garrison, T., 2019, Testing the efficacy of a Cyclops 7 infrared probe by performing a follow up dye trace study at McConnell Springs in Lexington, Kentucky, USA: Professional Geologist, Vol. 56, No. 4, pp. 50–54. Hess, J. W. and White, W. B., 1988, Storm response of the karstic carbonate aquifer of southcentral Kentucky: Journal Hydrology, Vol. 99, pp. 235–252. Hibbs, B. J. and Sharp, J. M., Jr., 2012, Hydrological impacts of urbanization: Environmental Engineering Geoscience, Vol. 18, No. 1, pp. 3–24. Husic, A.; Fox, J.; Adams, E.; Backus, J.; Pollock, E.; Ford, W.; and Agouridis, C., 2019, Inland impacts of atmospheric river and tropical cyclone extremes on nitrate transport and stable isotope measurements: Environmental Earth Sciences, Vol. 78, Article 36, https://doi.org/10.1007/s12665-018-8018-x Jeelani, G.; Shah, R. A.; Deshpande, R. D.; Fryar, A. E.; Perrin, J.; and Mukherjee, A., 2017, Distinguishing and estimating recharge to karst springs in snow and glacier dominated mountainous basins of the western Himalaya, India: Journal Hydrology, Vol. 550, pp. 239–252. Kendall, C. and Coplen, T. B., 2001, Distribution of oxygen-18 and deuterium in river waters across the United States: Hydrological Processes, Vol. 15, pp. 1363–1393. Kentucky American Water Company, 2022, 2021 Annual Water Quality Report, Central Division: Electronic document, available at https://www.amwater.com/ccr/lexington.pdf Kentucky Geological Survey, 2022, Kentucky Geologic Map Service: Electronic document, available at https:// kgs.uky.edu/kygeode/geomap KSIGL (Kentucky Stable Isotope Geochemistry Laboratory), 2022, δ2 H and δ18 O in Liquids: Electronic document, available at https://isotopes.as.uky.edu/liquids
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
Hydrochemical Delineation of Spring Recharge in an Urbanized Karst Basin, Central Kentucky Lakey, B. and Krothe, N. C., 1996, Stable isotopic variation of storm discharge from a perennial karst spring, Indiana: Water Resources Research, Vol. 32, pp. 721–731. Lewis, C. K., 2018, personal communication, Department of Environmental Quality Management, University of Kentucky, Lexington, KY. LFUCG (Lexington-Fayette Urban County Government), 2022, Snow and Ice Control Plan: Electronic document, available at https://www.lexingtonky.gov/snow-and-ice-controlplan Lim, K. J.; Engel, B. A.; Tang, Z.; Choi, J.; Kim, K. S.; Muthukrishnan, S.; and Tripathy, D., 2005, Automated Web GIS Based Hydrograph Analysis Tool, WHAT: Journal American Water Resources Association, Vol. 41, pp. 1407–1416. Liu, Z. H.; Groves, C.; Yuan, D. X.; Meiman, J.; Jiang, G. H.; He, S. Y.; and Li, Q. A., 2004, Hydrochemical variations during flood pulses in the south-west China peak cluster karst: Impacts of CaCO3 -H2 O-CO2 interactions: Hydrological Processes, Vol. 18, pp. 2423–2437. Luhmann, A. J.; Covington, M. D.; Alexander, S. C.; Chai, S. Y.; Schwartz, B. F.; Groten, J. T.; and Alexander, E. C., Jr., 2012, Comparing conservative and nonconservative tracers in karst and using them to estimate flow path geometry. Journal Hydrology, Vol. 448–449, pp. 201–211. Luhmann, A. J.; Covington, M. D.; Peters, A. J.; Alexander, S. C.; Anger, C. T.; Green, J. A.; Runkel, A. C.; and Alexander, E. C., Jr., 2011, Classification of thermal patterns at karst springs and cave streams: Groundwater, Vol. 49, pp. 324–335. Luo, M.; Chen, Z.; Zhou, H.; Zhang, L.; and Han, Z., 2018, Hydrological response and thermal effect of karst springs linked to aquifer geometry and recharge processes: Hydrogeology Journal, Vol. 26, pp. 629–639. Marx, C.; Tetzlaff, D.; Hinkelmann, R.; and Soulsby, C., 2021, Isotope hydrology and water sources in a heavily urbanized stream: Hydrological Processes, Vol. 35, e14377, https://doi.org/10.1002/hyp.14377 Mehlhorn, D. K., 2016, Project Final Report for Wolf Run Creek Watershed, Water Quality Best Management Practices, Picadome Golf Course: Electronic document, available at https://eec.ky.gov/Environmental-Protection/Water/ Reports/Reports/NPS1205-WolfRun-Picadome.pdf MRCC (Midwest Regional Climate Center), 2022, cli-MATE: MRCC Application Tools Environment: Electronic document, available at https://mrcc.purdue.edu/CLIMATE MRLCC (Multi-Resolution Land Characteristics Consortium), 2022, NLCD 2011 Land Cover (CONUS): Electronic document, available at https://www.mrlc.gov/data/nlcd2011-land-cover-conus Mudarra, M.; Andreo, B.; Marín, A. I.; Vadillo, I.; and Barbará, J. A., 2014, Combined use of natural and artificial tracers to determine the hydrogeological functioning of a karst aquifer: The Villanueva del Rosario system (Andalusia, southern Spain): Hydrogeology Journal, Vol. 22, pp. 1027–1039. National Institute of Standards and Technology, 2022, Engineering Statistics Handbook, 7.1.6, What Are Outliers in the Data?: Electronic document, available at https://www.itl.nist. gov/div898/handbook/prc/section1/prc16.htm Ozyurt, N. N. and Bayari, C. S., 2007, Temporal variation of chemical isotopic signals in major discharges of an alpine karst aquifer in Turkey: Implications with respect to response of karst aquifers to recharge: Hydrogeology Journal, Vol. 16, pp. 297–309.
Passarello, M. C.; Sharp, J. M., Jr.; and Pierce, S. A., 2012, Estimating urban-induced artificial recharge: A case study for Austin, TX: Environmental Engineering Geoscience, Vol. 18, No. 1, pp. 25–36. Phillips, J. D.; Martin, L. L.; Nordberg, V. G.; and Andrews, W. A., Jr., 2004, Divergent evolution in fluviokarst landscapes of central Kentucky: Earth Surface Processes Landforms, Vol. 29, pp. 799–819. Price, D. J., 2022, Over a decade of monitoring water quality at McConnell Springs—What have we learned? In Kentucky Water Resources Annual Symposium: Kentucky Water Resources Research Institute, University of Kentucky, Lexington, KY, pp. 10–11. PV and Associates, 2019, WinSLAMM: Electronic document, available at http://winslamm.com R Core Team, 2019, R: A Language and Environment for Statistical Computing, version 3.3.1. R Foundation for Statistical Computing, Vienna, Austria. Reisch, C. E. and Toran, L., 2014, Characterizing snowmelt anomalies in hydrochemographs of a karst spring, Cumberland Valley, Pennsylvania (USA): Evidence for multiple recharge pathways: Environmental Earth Sciences, Vol. 72, pp. 47–58. Robinson, H. K. and Hasenmueller, E. A., 2017, Transport of road salt contamination in karst aquifers and soils over multiple timescales: Science Total Environment, Vol. 603–604, pp. 94–108. Sharp, J. M., Jr., 2019, Effects of urbanization on the Edwards Aquifer. In Sharp, J. M., Jr.; Green, R. T.; and Schindel, G. M. (Editors), The Edwards Aquifer: The Past, Present, and Future of a Vital Water Resource: Memoir 215, Geological Society of America, Boulder, CO, pp. 213–222. Smith, D. F.; Saelens, E.; Leslie, D. L.; and Carey, A. E., 2021, Local meteoric water lines describe extratropical precipitation: Hydrological Processes, Vol. 35, No. e14059. Soil Survey Staff, 2023a, Gridded Soil Survey Geographic (gSSURGO) Database for the Conterminous United States: Electronic document, available at https://gdg.sc.egov.usda. gov Soil Survey Staff, 2023b. Web Soil Survey: Electronic document, available at https://websoilsurvey.sc.egov.usda.gov/app Soulsby, C.; Birkel, C.; and Tetzlaff, D., 2014, Assessing urbanization impacts on catchment transit times, Geophysical Research Letters, Vol. 41, pp. 442–448. Stevenson, J. L.; Geris, J.; Birkel, C.; Tetzlaff, D.; and Soulsby, C., 2022, Assessing land use influences on isotopic variability and stream water ages in urbanising rural catchments: Isotopes Environmental Health Studies, Vol. 58, pp. 277–300. Stroud Water Research Center, 2022, Model My Watershed Site Storm Model: Electronic document, available at: https://modelmywatershed.org Tian, C.; Wang, L.; Kaseke, K. F.; and Bird, B. W., 2018, Stable isotope compositions (δ2 H, δ18 O and δ17 O) of rainfall and snowfall in the central United States: Scientific Reports, Vol. 8, No. 6712. Toran, L.; Gross, K.; and Yang, Y., 2009, Effects of restricted recharge in an urban karst system: Environmental Geology, Vol. 58, pp. 131–139. USGS (U.S. Geological Survey), 2022, Wolf Run at Old Frankfort Pike at Lexington, KY: Electronic document, available at https://waterdata.usgs.gov/monitoring-location/03289193 WHAT, 2022, WHAT: Web-based Hydrograph Analysis Tool: Electronic document, available at https://engineering. purdue.edu/mapserve/WHAT
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
215
Fryar, Currens, and Alvarez Villa White, W. B., 2002, Karst hydrology: Recent developments and open questions: Engineering Geology, Vol. 65, pp. 85–105. Zhang, Z.; Chen, X.; Cheng, Q.; Li, S.; Yue, F.; Peng, T.; Waldron, S.; Oliver, D. M.; and Soulsby, C., 2020, Coupled hydrological and biogeochemical modelling of nitrogen transport in the karst critical zone:
216
Science Total Environment, Vol. 732, Article 138902, https://doi.org/10.1016/j.scitotenv.2020.138902 Zhang, Z.; Chen, X.; Cheng, Q.; and Soulsby, C., 2019, Storage dynamics, hydrological connectivity and flux ages in a karst catchment: Conceptual modelling using stable isotopes: Hydrology Earth System Sciences, Vol. 23, pp. 51–71.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 203–216
Open Access Article
Assessing Landscape and Seasonal Controls on Soil CO2 Fluxes in a Karst Sinkhole TARYN K. THOMPSON School of Plant and Environmental Sciences, Virginia Tech, Blacksburg, VA 24061
DANIEL L. McLAUGHLIN Department of Forest Resources and Environmental Conservation, Virginia Tech, Blacksburg, VA 24061
MADELINE E. SCHREIBER Department of Geosciences, Virginia Tech, Blacksburg, VA 24061
RYAN D. STEWART* School of Plant and Environmental Sciences, Virginia Tech, Blacksburg, VA 24061
Key Terms: Diffusion, Carbon Cycling, Gas Exchange, Mass Transport, Zero-Flux Plane ABSTRACT Carbon dioxide (CO2) gas diffusion is an important component of carbon cycling in soils. This process is particularly relevant in karst landscapes, which contain easily weathered rock, subsurface fractures, and cave networks. We instrumented three soil profiles— the shoulder, back slope, and toe slope of a sinkhole— above a karst cave in Virginia. Each profile had solidstate CO2 sensors and soil water content/temperature sensors at 20 and 60 cm depth that collected hourly measurements from 2017 to 2019. We calculated CO2 fluxes using Fick’s first law along with measured soil and assumed atmospheric CO2 concentrations. With this approach, we identified occasional near-surface zero-flux planes, in which CO2 likely diffused both upward and downward. All profiles had upward CO2 fluxes during warm-season months, with maximum fluxes of 1.2 lmol CO2 m22 s21 in the shoulder and back slope versus 2.0 lmol CO2 m22 s21 in the toe slope. During cool-season months, upward CO2 fluxes were smaller (0–0.3 lmol CO2 m22 s21) and were often counteracted by downward fluxes in the toe slope, possibly driven by ventilation into the underlying cave. The toe slope had a cumulative annual efflux of 14.5 mol CO2 m22, which was .3 times greater than the other profiles. Fluxes were sensitive to soil porosity, with an order-of-magnitude difference when porosity was assumed *Corresponding author email: rds@vt.edu
to be 0.40 versus 0.56 cm3 cm23. The results of this study offer new insight into short-term and seasonal variations in diffusive CO2 gas transport in karst soils, and they may inform other investigations of non-uniform diffusion processes. INTRODUCTION Exchange of carbon dioxide (CO2) from the land surface to the atmosphere is an important process in the global carbon cycle, yet the controls and magnitudes of CO2 emissions from the soil surface remain poorly constrained (Jian et al., 2022). Carbon dynamics are of particular interest in karst regions, which are characterized by relatively soluble carbonate rock, usually dolomite or limestone rock, and which occupy as much as 15 percent of the global land surface (Ford and Williams, 2007). Over timescales of thousands to hundreds of thousands of years, karst regions absorb CO2 during chemical weathering of carbonate rocks. Much of the CO2 involved in this weathering process comes from overlying soils or the atmosphere, with gas diffusion found to be a more important process than off-gassing from aqueous solutions (Breecker et al., 2012). Thus, understanding the magnitudes and directions of CO2 fluxes in karst soils is necessary to perform accurate regional and global carbon budgeting. Subsurface transport of gases such as CO2 occurs via two processes, diffusion and advection. Advection occurs as a mass flow in which gases are typically transported via flowing wind or water. Diffusion occurs due to concentration gradients driving gas movement from areas of high to low concentration, and it is the more important process for gas exchange in most soils (Pumpanen et al.,
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
217
Thompson, McLaughlin, Schreiber, and Stewart
2003; Laemmel et al., 2017). The rate of gas diffusion depends on two factors: (1) the diffusion coefficient, Ds, and (2) the concentration gradient, (i.e., DC/Dx for one-dimensional transport). The diffusion coefficient is a measure of the rate of movement of a gas under constant temperature and pressure conditions through a uniform medium. In soils, Ds is affected by several factors, with the volume of air-filled pore space representing a particularly important term (Jin and Jury, 1996; Lafond et al., 2011). Concentration gradients can be quantified by measuring partial pressures, or equivalent molar concentrations, at multiple depths. In typical soils, CO2 concentrations tend to increase with depth due to respiration and absence of any large-scale sinks (Tang et al., 2003). Conversely, studies conducted in karst areas have identified conditions under which soil CO2 concentrations near the surface are greater than those at depth (Chen, 2019). One possibility for these observations is CO2 consumption during dissolution of carbonate minerals (Chen, 2019; Zhao et al., 2021). A second possibility is that venting of cave air removes CO2 from deeper soil and epikarst; for example, Benavente et al. (2010) found that 30–50 percent of CO2 in a Mediterranean cave originated from the overlying soil profile, suggesting downward diffusion of CO2 from soils into the cave. This venting process typically occurs during cooler seasons of the year, when air inside the cave system is warmer and less dense than the surrounding air (Mattey et al., 2016; Krajnc et al., 2017; Riechelmann et al., 2019). These observations imply that a zero-flux CO2 plane can exist in the subsurface, from which gas diffuses both upward toward the soil surface and downward through the subsurface, similar in concept to hydraulic zero-flux planes that form in soils (Brutsaert, 2014). However, limited research has been conducted to assess the frequency with which near-surface zero-flux planes exist, and the corresponding implications for gas diffusion remain poorly understood. Variation in wetness, soil depth, and other environmental factors associated with topography can also affect CO2 fluxes. Previous work has shown that riparian areas often have greater soil respiration than surrounding uplands. This result is likely due to wetter conditions encouraging greater ecosystem productivity (Webster et al., 2008; Pacific et al., 2009, 2011), which can supersede smaller Ds values found in riparian compared to upland soils (Pacific et al., 2008). Other work focused on hillslope profiles has shown that ridge tops often have greater soil respiration rates than lower-lying positions (Tian et al., 2019), particularly in wet years (Kopp et al., 2022). Slope angle can also affect soil respiration and carbon allocation, with flatter or concave slopes found to have greater respiration than steeper or convex slopes (Saggar et al., 1999; Fissore et al., 2017). It is likewise probable that topography affects soil respiration and gas diffusion in karst areas, but the controls on CO2 gas movement in such systems have not yet been critically examined. 218
To evaluate the magnitudes and directions of CO2 fluxes in a karst landscape, we instrumented three locations along a sinkhole slope—shoulder, back slope, and toe slope—and measured hourly CO2 concentrations at 20 and 60 cm depths from winter 2017 to summer 2019. We used those measurements, along with assumed atmospheric CO2 concentration and data for soil physical properties (e.g., porosity, water content, temperature), to calculate CO2 fluxes. Our first objective was to compare upward and downward CO2 fluxes for the three sinkhole positions. We hypothesized that the upward fluxes would be similar in magnitude between the three soil profiles, but the downward fluxes would be largest in the toe slope and smallest in the shoulder profile, based on differences in vertical distance from the underlying cave system. Our second objective was to examine seasonal trends in CO2 fluxes. Here, we hypothesized that CO2 fluxes would be smaller during cooler months of the year, due to limited soil respiration along with wet conditions limiting gas exchange, and fluxes would also include a sizeable downward component due to convective venting from the underlying cave system. Conversely, we hypothesized that CO2 fluxes would be larger and primarily upward during warmer months of the year, because respiration typically increases with temperature, and gas diffusion rates typically increase with lower soil water contents (Fang and Moncrieff, 2005; Ali et al., 2018). Our third objective was to evaluate the sensitivity of gas fluxes to model parameters used to estimate gas diffusivity. We expected that the estimated flux would be sensitive to available pore space in the soil profiles (i.e., the difference between soil porosity and soil water content), and that the effects would be strongest in the non-growing season, when water contents are typically high, compared to the growing season. METHODS Site Description The field site was located at James Cave in southwest Virginia (Supplemental Material Figure S1, https:// www.aegweb.org/e-eg-supplements). James Cave is a doline karst cave system with substantial amounts of limestone and dolomite rock (Eagle et al., 2015; Schreiber et al., 2015). Based on the U.S. Department of Agriculture Soil Survey (Soil Survey Staff, 2019), the soils surrounding James Cave are classified as Slabtown silt loam, Lowell silt loam, and Wurno-Newbern-Faywood silt loam, with soil thickness ranging from 0.5 to 2.0 m (Eagle et al., 2015). The field site was managed as a beef cattle pasture, similar to the predominant land use of the surrounding area. The vegetation was a mixed stand of cool-season grasses, including tall fescue (Festuca arundinacea).
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
Controls on CO2 Fluxes in Karst 10 m
12 m
Shoulder 3m
Backslope 2m
Toeslope
James Cave
Figure 1. (Top) Conceptual drawing of the study site (James Cave, VA) including approximate horizontal and vertical dimensions between the sinkhole slope locations and associated sensor depths relative to the sinkhole slope and James Cave entrance. (Bottom left) Site photo indicating the entrance to James Cave and the locations of the three instrumented profiles. (Bottom right) James Cave extent and surface topography, with locations of the instrumented profiles used in this study indicated by red X’s. Cave survey is courtesy of Tom Malabad.
Sensors were installed in December 2016 in the sinkhole that surrounds the entrance to James Cave, with vertical profiles instrumented at the shoulder, back slope, and toe slope of the sinkhole slope (Figure 1). Each profile contained two CO2 sensors (Vaisala GMM221, 5 percent maximum concentration, Vaisala Corporation, Vantaa, Finland) and two soil water content and temperature sensors (CS655, Campbell Scientific, Logan, UT) located at 20 and 60 cm below the surface. The 60 cm soil water content/temperature sensor did not properly function at the back-slope location, so an additional sensor was installed at that depth (5TM, Decagon Devices, Pullman, WA). The data were measured every hour using a CR1000 logger (Campbell Scientific, Logan, UT). All installed sensors functioned from 7 February 2017 through 8 July 2018, when the 20-cm-deep CO2 sensor failed at the back slope. The 60-cm-deep CO2 sensor failed at the shoulder location on 24 September 2018, and the 20-cm-deep CO2 sensors failed at the toe slope on 19 September 2019. All three profiles therefore had a common 1.5 year period in which both CO2 sensors functioned, and the toe slope had an additional year of data. The barometric pressure and hourly precipitation data were downloaded through the National Oceanic and Atmospheric Administration’s National Centers for Environmental Information from the Integrated Surface Database for the Blacksburg Airport station, which was located approximately 17 km from the study site.
Temperature readings came from the adjacent soil temperature sensors. Soil Characterization Volumetric soil cores were collected from the site in order to determine water retention properties, which we used to infer total porosity and macro- versus microporosity. The cores were collected from depths of 15, 20, 35, and 40 cm for the shoulder location and from 15, 20, and 40 cm depths for the back-slope and toe-slope locations. The cores were first saturated with 0.01 M CaCl2 solution and were then analyzed for water retention using a positive pressure system (SoilMoisture, Inc., Santa Barbara, CA). Soil water content was determined at pressure heads of 0 cm (saturation), 340 cm (assumed to represent field capacity), 1,020 cm, and 3,060 cm. Porosity was calculated as the water content at saturation, macro-porosity was calculated as the difference in water content between saturation and 340 cm, and micro-porosity was calculated as the water content at 340 cm. CO2 Flux Calculations The measured CO2 concentrations were corrected for temperature and pressure based on the ideal gas law (Vaisala, 2017):
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
219
Thompson, McLaughlin, Schreiber, and Stewart
Ccorrected ¼ Cmeasured
Pstd ; RðTstd þ 5Þ
(1)
where Ccorrected is the CO2 concentration corrected for temperature and pressure [N N 1], Cmeasured is the measured CO2 concentration [N N 1], Pstd is the standard barometric pressure used in the instrument calibration (101.3 kPa), Tstd is the standard temperature (293 K), and R is the ideal gas law constant (8.31 3 10 3 kPa m3 mol 1 K 1). We next calculated CO2 fluxes, J [N L 2 t 1], using Fick’s first law: J ¼ Ds
DC ; Dx
(2)
where DC is the difference in CO2 concentrations [N N 1] for two locations separated by a vertical distance Dx [L], and Ds is the diffusion coefficient [L2 t 1]. Ds was calculated following Tang et al. (2003): ! e10=3 Ds ¼ D0 ; (3) /2 e ¼ / h; D0 ¼ Da0
T Tstd
(4)
1:75
P ; Pstd
3
3
(5)
where e is the air-filled porosity [L L ], / is the soil porosity [L3 L 3], h is the soil water content [L3 L 3], T is temperature [K], P is air pressure [M L 1 t 2], and Da0 is the gas diffusivity in pure air under standard temperature and pressure conditions [L2 t 1], assumed here as Da0 ¼ 1.4 3 10 5 m2 s 1 (Pritchard and Currie, 1961). Porosity, /, was determined based on the mean volumetric saturated water content (hs) from the soil cores collected in each profile. For every hourly time step, we calculated two fluxes: one flux representing the upper 20 cm (J0–20) that used the measured Ccorrected value at 20 cm and an assumed atmospheric CO2 concentration of 400 lmol mol 1, and a second flux representing the 20–60 cm depth interval (J20–60) that used Ccorrected values from those two depths. To calculate the gas diffusivity constant for J0–20, we used measured T and h values from 20 cm, and for J20–60, we used the mean T and h values from the 20 and 60 cm sensors. The measured Ccorrected values at 20 cm were always greater than the assumed atmospheric concentration. Therefore, whenever Ccorrected at 20 cm was greater than Ccorrected at 60 cm, we assumed that a near-surface zero-flux existed for CO2 diffusion, and we calculated an upward flux in the upper 20 cm of the soil profile (i.e., Jupward ¼ J0–20) and a downward flux from 20 to 60 cm (i.e., Jdownward ¼ J20–60). Whenever 220
Ccorrected at 20 cm was less than or equal to Ccorrected at 60 cm, we calculated only an upward flux using a depth-weighted average of the two fluxes (i.e., Jupward ¼ 1/3 3 J0–20 þ 2/3 3 J20–60). This approach assumed isothermal conditions within the soil profile (Jones, 2014). There were some brief periods in which the CO2 concentrations at one or more locations were out of sensor range (i.e., .55,000 ppmv or 5.5 percent); fluxes were not estimated for those times. We also quantified seasonal and annual CO2 fluxes from each profile, with the upward and downward fluxes assessed separately for each profile. Annual fluxes were calculated by summing hourly fluxes between 1 April 2017 and 31 March 2018. For the seasonal fluxes, we divided the data into the warm, growing season from 1 April 2017 to 30 September 2017, and the cool, nongrowing season from 1 October 2017 to 31 March 2018. We used linear interpolation to fill gaps in the hourly flux records, using the flux values that were calculated immediately before and after the missing records. This approach gave similar values as using only records with measured fluxes (data not shown), indicating that the linear interpolation process did not overly influence the results. Sensitivity Analysis We performed a brief sensitivity analysis to test how the porosity (/) term in Eq. 3 and Eq. 4 affected calculations of Ds and Jupward. First, we calculated Ds through time for each profile using five assumed porosities: / ¼ 0.40, 0.44, 0.48, 0.52, and 0.56 cm3 cm 3. These porosities reflected the range of values measured in the soil samples used for soil characterization. The Ds term was assumed to equal 0 any time that the measured volumetric water content was equal to or greater than the assumed porosity. We next used Eq. 2 to Eq. 5 to determine upward CO2 fluxes for each profile and then summed the cumulative CO2 efflux for the period between 1 April 2017 and 31 March 2018. RESULTS Soil Properties The water retention curves were similar between most soil cores (Supplemental Material Figure S2, https:// www.aegweb.org/e-eg-supplements). For example, at field capacity (i.e., pressure head ¼ 340 cm), volumetric water content values were between 0.32 and 0.41 cm3 cm 3 for the toe-slope location, 0.33 and 0.39 cm3 cm 3 for the back-slope location, and 0.33 and 0.38 cm3 cm 3 for the shoulder location. Similar ranges of water contents were seen for the 1,020 cm and 3,060 cm pressure heads.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
Controls on CO2 Fluxes in Karst 2.0
Shoulder Sho oulde er
Backslope Backslo ope
Toeslope Toeslo ope
Ds (mm2 s-1)
1.5
1.0
0.5
0.0 Jan-2017 May-2017 Sep-2017 Jan-2018 May-2018 Sep-2018 Jan-2019 May-2019 Sep-2019
Figure 2. Gas diffusivity, Ds, values calculated for the three soil profiles (shoulder, back slope, and toe slope) using Eqs. 3–5 and mean measured soil water contents and temperatures from 20 and 60 cm depths. Porosity values were estimated using mean values determined from soil cores collected in each profile. The cool seasons from October through March are indicated by the blue-shaded regions, and the warm seasons from April through September are indicated by the yellow shading.
We used these water retention values to calculate total, macro-, and micro-porosities (Supplemental Material Table S1, https://www.aegweb.org/e-eg-supplements). Porosity varied between 0.41 and 0.55 cm3 cm 3 for the toe-slope location, 0.41 and 0.49 cm3 cm 3 for the back-slope location, and 0.40 and 0.47 cm3 cm 3 for the shoulder location. We therefore used mean values of / ¼ 0.48 cm3 cm 3 for the toe-slope profile, / ¼ 0.45 cm3 cm 3 for the backslope profile, and / ¼ 0.44 cm3 cm 3 for the shoulder when calculating Ds. The macro-porosities varied from 0.06 to 0.15 cm3 cm 3, with the toe-slope location having greater macro-porosity at 15–20 cm compared to the other two profiles. Micro-porosity varied from 0.30 to 0.40 cm3 cm 3, with similar sample-to-sample variation in all three profiles. Environmental Conditions The largest precipitation events occurred in the summer and early fall, with maximum rates of 1.0 to 2.5 cm hr 1 (Supplemental Material Figure S3a, https://www.aegweb.org/ e-eg-supplements). Volumetric water contents were lower in the summer and fall months, with a range of 0.10 to 0.35 cm3 cm 3. During the cool season (i.e., the months of October through March), the volumetric water contents were higher, varying between 0.30 and 0.45 cm3 cm 3. The shoulder and toe-slope locations tended to experience the greatest seasonal differences in volumetric water contents, while the back slope had smaller-magnitude changes throughout the study period. Soil temperatures also fluctuated seasonally, varying between 15°C and 26°C for the warmest months (June– September) and between 1°C and 13°C in the coolest months (December–March; Supplemental Material Figure S3b, https://www.aegweb.org/e-eg-supplements). Soil temperatures showed marked transitions between the colder and
warmer extremes during early spring (i.e., March through May) and early fall (i.e., September through November). Gas Diffusivity Constant and CO2 Fluxes Calculated values for gas diffusivity, Ds, varied through time due to differences in soil temperatures and water content, and also varied among the three soil profiles due to those factors and different assumed values for / (Figure 2). In the first year of monitoring, the toe slope had peak Ds values that were nearly three times the values of the other two profiles. In the second and third years, the shoulder and toe slope profiles had similar Ds values, while the back slope continued to have smaller values. In all three profiles, the Ds values were larger during the warm season and were often near 0 during the cool seasons. The shoulder profile had relatively high CO2 concentrations early in the warm season, at times exceeding 55,000 ppmv (5.5 percent), and had relatively low CO2 concentrations during the cool, non-growing season (Figure 3a). The 60 cm sensor tended to have higher concentrations than the 20 cm sensor, though there were many brief periods in which the concentration gradient became inversed (i.e., greater CO2 concentration at 20 cm than 60 cm). Those periods were indicative of a near-surface zero-flux plane, and they translated to relatively large upward CO2 fluxes and much smaller downward fluxes (Figure 3b). During the cool seasons, the profile had negligible fluxes in either direction. Upward CO2 fluxes peaked between 0.5 and 1.2 lmol CO2 m 2 s 1, while the magnitude of downward fluxes never exceeded 0.2 lmol CO2 m 2 s 1. The back-slope profile had similar magnitudes and seasonal trends in CO2 concentrations as the shoulder, with CO2 concentrations peaking above 55,000 ppmv during the beginning of the warm season and then
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
221
Thompson, McLaughlin, Schreiber, and Stewart a)
60000
CO2 concentration (μmol mol-1)
20 cm 50000
60 cm
40000
30000
20000
10000
0 Dec-2016 Apr-2017 Aug-2017 Dec-2017 Apr-2018 Aug-2018 Dec-2018 Apr-2019 Aug-2019
CO2 flux (μmol m-2 s-1)
b)
2.0
1.5
1.0
0.5
0.0
-0.5
-1.0 Dec-2016 Apr-2017 Aug-2017 Dec-2017 Apr-2018 Aug-2018 Dec-2018 Apr-2019 Aug-2019
Figure 3. (a) CO2 concentrations at 20 and 60 cm depths, and (b) upward (black) versus downward (red) CO2 fluxes in the shoulder location. The cool seasons from October through March are indicated by the blue-shaded regions, and the warm seasons from April through September are indicated by yellow shading.
gradually declining (Figure 4a). The CO2 fluxes for the back-slope location had similar magnitudes of upward fluxes as the shoulder location (Figure 4b). However, during the first warm season, the downward fluxes were consistently larger in magnitude than those detected in the shoulder profile. Furthermore, the downward and upward fluxes often had similar magnitudes in this profile, with larger fluxes in both directions during warm seasons and near-zero fluxes during the cool seasons. The toe-slope location had more consistently elevated CO2 concentrations compared to the other two profiles (Figure 5a). Here, the 20 cm sensor often recorded greater CO2 concentrations than the 60 cm sensor, particularly during the cool, non-growing season. The toe-slope profile also tended to have larger CO2 fluxes than the other two profiles (Figure 5b). The CO2 fluxes reached peak values of 1.5–2.0 lmol CO2 m 2 s 1 (Jupward) and 0.5 lmol CO2 m 2 s 1 (Jdownward). Downward fluxes were often present during 222
the cool season, with larger magnitudes during the winter of 2017–2018 as compared to 2018–2019. Annual CO2 efflux for each profile was calculated by integrating the areas underneath the CO2 flux curves for each profile from 1 April 2017 through 31 March 2018. The annual efflux values ranged from 3.5 mol CO2 m 2 in the shoulder to 14.5 mol CO2 m 2 in the toe slope (Table 1). Most of the efflux occurred during the growing season, with warm-season efflux representing between 76 percent (in the toe slope) and 84 percent (in the back slope) of total CO2 emissions. The cumulative downward flux was much smaller, with the toe slope having a cumulative total of 0.36 mol CO2 m 2, the back slope having a cumulative total of 0.13 mol CO2 m 2, and the shoulder having a cumulative total of 0.04 mol CO2 m 2. Sensitivity Analysis The sensitivity analysis showed that the magnitude of Ds varied widely based on the assumed value of /
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
Controls on CO2 Fluxes in Karst
CO2 concentration (μmol mol-1)
a)
60000
20 cm 50000
60 cm 40000
30000
20000
10000
0 Dec-2016 Apr-2017 Aug-2017 Dec-2017 Apr-2018 Aug-2018 Dec-2018 Apr-2019 Aug-2019
b)
2.0
CO2 flux (μmol m-2 s-1)
1.5
1.0
0.5
0.0
-0.5
-1.0 Dec-2016 Apr-2017 Aug-2017 Dec-2017 Apr-2018 Aug-2018 Dec-2018 Apr-2019 Aug-2019
Figure 4. (a) CO2 concentrations at 20 and 60 cm depths, and (b) upward (black) versus downward (red) CO2 fluxes in the back-slope location. The cool seasons from October through March are indicated by the blue-shaded regions, and the warm seasons from April through September are indicated by yellow shading.
(Figure 6). Here, we use the toe-slope location as an example. Under the lowest tested / value (0.40 cm3 cm 3), the maximum Ds values, based on mean h and T values from the 20 and 60 cm sensors, were between 1.0 and 1.2 mm2 s 1. The largest tested / value (0.56 cm3 cm 3) gave a maximum Ds value of 2.8 mm2 s 1. In the cool season, when the soil water content was high, the variation in porosity drove the Ds value closer to zero for the minimum porosity value versus 0.1–0.2 mm2 s 1 when larger porosity values were used. The differences in calculated Ds values translated to sizeable differences in the estimated amount of CO2 that was emitted from each soil profile over a 1 year period (Figure 7). The toe-slope location had the least variation, with total CO2 efflux varying by 73 between the lowest (/ ¼ 0.40) and highest (/ ¼ 0.56) porosities. Cumulative effluxes between lowest and highest porosities differed by 113 for the shoulder profile and by 143 in the back-slope location. These
results show the importance of accurately constraining porosity when estimating CO2 fluxes using solid-state sensors embedded in the soil. DISCUSSION CO2 Fluxes Increased during the Growing Season and Were Larger in the Sinkhole Toe Slope We initially hypothesized that the upward CO2 fluxes would be similar in magnitude for the three profiles, whereas the shoulder would have the smallest downward CO2 fluxes, and the toe slope would have the largest downward fluxes. We also hypothesized that upward CO2 fluxes would be much greater in the warm season than the cold season, whereas downward fluxes would be dominant in the cold season due to convective venting from the cave releasing CO2 back to the atmosphere. The study results provided support for these hypotheses,
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
223
Thompson, McLaughlin, Schreiber, and Stewart a) CO2 concentration (μmol mol-1)
60000
20 cm
50000
60 cm 40000
30000
20000
10000
0 Dec-2016 Apr-2017 Aug-2017 Dec-2017 Apr-2018 Aug-2018 Dec-2018 Apr-2019 Aug-2019
b)
2.0
CO2 flux (μmol m-2 s-1)
1.5
1.0
0.5
0.0
-0.5
-1.0 Dec-2016 Apr-2017 Aug-2017 Dec-2017 Apr-2018 Aug-2018 Dec-2018 Apr-2019 Aug-2019
Figure 5. (a) CO2 concentrations at 20 and 60 cm depths, and (b) upward (black) versus downward (red) CO2 fluxes in the toe-slope location. The cool seasons from October through March are indicated by the blue-shaded regions, and the warm seasons from April through September are indicated by yellow shading.
though some discrepancies and nuances emerged during our analysis. The back-slope and shoulder profiles had similar upward CO2 fluxes, partially supporting our hypothesis. Table 1. Calculated cumulative CO2 efflux for the three sinkhole slope positions over different time periods. The annual period included 1 April 2017 to 31 March 2018, the warm season included 1 April 2017 to 30 September 2017, and the cool season included 1 October 2017 to 31 March 2018. Back Slope
Toe Slope
Cumulative Upward Flux (mol CO2 m 2) Annual 365 3.54 Warm Season 183 2.90 Cool Season 182 0.64
No. of Days
4.46 3.75 0.71
14.5 11.1 3.43
Cumulative Downward Flux (mol CO2 m 2) Annual 365 0.04 Warm Season 183 0.03 Cool Season 182 0.01
0.13 0.11 0.02
0.36 0.09 0.27
224
Shoulder
The toe-slope profile, however, had more sustained and often greater upward fluxes, with peak fluxes near twice those of the other two profiles (i.e., 2.0 versus 1.0– 1.2 lmol m 2 s 1). These differences primarily occurred during the warm growing season of the first study year, when the shoulder and back-slope profiles had smaller Ds estimates than the toe slope. In the second warm season, by comparison, the shoulder and toe slope had similar Ds values, and the two profiles had similar peak effluxes (1.3–1.6 lmol m 2 s 1). These comparisons indicate that CO2 production was likely similar between locations, and the differences in efflux were attributable to variations in soil moisture conditions and resulting estimates for air-filled porosity e. All three profiles also had periods in which the shallow (20 cm) sensors recorded higher CO2 concentrations compared to their deeper (60 cm) counterparts, implying the existence of a zero-flux CO2 plane in the shallow subsurface. Downward diffusion likely occurred whenever that condition existed, and annual total downward fluxes were
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
Controls on CO2 Fluxes in Karst 3.0
φ = 0.40
φ = 0.4 0.44 44
φ = 0.48
φ = 0.52 0.52
φ=
2.5
Ds (mm2 s-1)
2.0
1.5
1.0
0.5
0.0 Jan-2017 May-2017 Sep-2017 Jan-2018 May-2018 Sep-2018 Jan-2019 May-2019 Sep-2019
Figure 6. Calculated gas diffusivity, Ds, for the toe-slope location assuming five different porosity values: / ¼ 0.40, 0.44, 0.48, 0.52, and 0.56 cm3 cm 3. The cool seasons from October through March are indicated by the blue-shaded regions, and the warm seasons from April through September are indicated by the yellow shading.
estimated to be 0.04–0.36 mol CO2 m 2. While considerably smaller in magnitude than cumulative upward fluxes, which ranged from 3.6 to 14.5 mol CO2 m 2, these results show that downward CO2 diffusion may be a non-negligible process. Comparing sinkhole locations, the toe-slope profile had the largest peak and cumulative downward fluxes, as we hypothesized. These fluxes occurred primarily in the cool, non-growing season, which also supported our hypothesis. However, downward fluxes in the shoulder and back-slope locations were detected primarily during warm months, which we had not expected. These contrasting results imply that the degree and timing of downward gas diffusion in karst soils may be governed by complex interactions among CO2 production, CO2 consumption during carbonate dissolution, subsurface properties, and landscape position. In temperate climates, cave ventilation increases during the cool season, as temperature-driven differences in air density allow warmer cave air to escape via cave openings and become replaced by atmospheric air that is Cumulative CO2 Efflux (mol m-2)
35 30 25 20
Toeslope Backslope Shoulder
15 10 5 0 0.35
0.40
0.45
0.50
0.55
0.60
ϕ (cm3 cm-3)
Figure 7. Sensitivity analysis showing calculated annual CO2 efflux (in mol m 2) between 1 April 2017 and 31 March 2018 for different porosity (/) values.
lower in CO2 (Meyer et al., 2014; Mattey et al., 2016). Under such conditions, elevated CO2 concentrations in the soil would result in downward diffusion, but only so long as Ds is sufficiently large. The toe-slope location, being closest in elevation to the cave system and likely to be underlain by fractures and fissures, could have maintained air-filled pathways between the soil profile and the deeper epikarst and cave layers (Weisbrod et al., 2009; Eagle et al., 2015). The water content measurements indicated that the toe-slope profile was rarely saturated, with sustained water contents of ,0.37 cm3 cm 3 during the winter period compared to measured porosities of 0.41–0.55 cm3 cm 3. Concurrent work examining cave drip patterns within James Cave also showed a prolonged dry period with limited cave recharge throughout much of the 2017–2018 cool season (Groce-Wright et al., 2022), suggesting that larger fissures and pores above the cave may have remained unsaturated. However, these air-filled pathways may not have been connected to the shoulder and back-slope profiles, which were located at a greater distance above the cave system and had near-saturated soil water contents throughout the cool season. The shoulder and back-slope profiles were instead found to only experience downward CO2 diffusion during dry periods of the year, when e may have been sufficient to enable rapid gas diffusion rates (Cuezva et al., 2011; Sanchez-Cañete et al., 2011). Besides cave ventilation, there are other potential mechanisms by which the underlying karst layers could serve as CO2 sinks. For one, carbon dioxide can dissolve into percolating water and become converted into carbonic acid (H2CO3) and then bicarbonate (HCO3 ) during carbonate rock weathering (Ford and Williams, 2007). This process could reduce CO2 concentrations at depth as carbonates undergo dissolution. Previous work in the James Cave system found carbonate weathering to be highly seasonal, with increased CO2 advection
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
225
Thompson, McLaughlin, Schreiber, and Stewart
identified during December to March, when aquifer recharge occurred via cave drips (Eagle et al., 2015). This period overlaps with the period of the most consistent downward fluxes in the toe slope, suggesting a possible linkage between these processes.
et al., 2020). Therefore, future work should endeavor to better understand how spatial differences in vadose zone properties—including /, soil macro-porosity, and epikarst fracture connectivity—influence Ds values and gas diffusion rates.
Sensitivity Analysis
Limitations of the Study
The CO2 fluxes calculated in this study were sensitive to the porosity (/) value used to calculate Ds, with the cumulative efflux calculated over a 1 year period being an order of magnitude less when using the smallest versus largest / values. The differences were most pronounced during the colder, non-growing season, when water contents were typically high. During those times, the calculated fluxes were negligible (i.e., near zero) for the lowest porosity values versus 0.1–0.3 lmol m 2 s 1 when the highest porosity of / ¼ 0.56 cm3 cm 3 was used. Previous work has shown that e is one of the most important terms for Ds (Jin and Jury, 1996; Lafond et al., 2011), as captured in the diffusion model used here (Eqs. 3–5). We based our assumed / values on measurements taken from soil cores, yet those measurements revealed substantial variability between cores (/ ranging from 0.40 to 0.55 cm3 cm 3). The actual / in the field also likely differed from the values measured from the cores, due to possible errors when collecting and analyzing the cores (e.g., incomplete or over-saturation) and due to any naturally occurring differences in porosity throughout the soil profiles, such as the near-surface compaction that can result from animal traffic (Daniel et al., 2002). Discrepancies in assumed versus actual porosity values may have the greater effect on flux estimates in wet conditions, when the presence of connected air-filled pathways becomes particularly important (Jin and Jury, 1996). Our approach also did not consider the possibility of temporarily reduced gas diffusion due to near-surface wetting events, which can cause over-estimates of CO2 efflux (Fan and Jones, 2014). These factors imply that the calculated CO2 fluxes in this study had considerable uncertainty, and they emphasize the importance of accurately constraining porosity when estimating CO2 fluxes using solid-state sensors embedded in the soil. Other measurements and models for Ds have identified the importance of larger macro-pores for gas diffusion (McCourty et al., 2018; Jayarathne et al., 2020). The soil samples had estimated macro-porosities of 0.06–0.15 cm3 cm 3 (Supplemental Material Table S1, https://www.aegweb.org/e-eg-supplements), with the toeslope location having a greater proportion of macro-pores compared to the other profiles. The larger pores also likely extended below the soil mantle, as highly weathered epikarst layers often contain many fractures and large pores (Eagle et al., 2015; Mattey et al., 2016; White, 2016; Cao
The study had several limitations. For one, many of the solid-state CO2 sensors reached maximum readings ( 55,000 ppmv) multiple times during the study, which resulted in data gaps. The sensors then started to fail after 1.5–2.5 years, affecting the duration of the observational record. In addition, the overall spacing between sensors in this study was large compared to other studies (e.g., 8 cm spacing in Tang et al., 2003), which increased the likelihood that the concentration gradients were not uniform between sensors. The shallowest sensor depth was 20 cm, which is substantially deeper than that used in many other studies, and this could have affected the accuracy of the calculated CO2 efflux to the atmosphere. We also did not directly measure CO2 concentrations in the cave, which limited our ability to fully understand the magnitude of downward CO2 diffusion. Other studies in similar cave systems have measured typical concentrations of ,12,000 ppmv (Breecker et al., 2012; Meyer et al., 2014; Kukuljan et al., 2021), albeit with isolated observations of CO2 concentrations up to 30,000 ppmv (Cowan et al., 2013). These values are, for the most part, much lower than the typical concentrations measured during the growing season in our soil profiles, which were often in excess of 25,000 ppmv. It is therefore possible that a zero-flux CO2 plane existed below our deepest sensor. If so, the instantaneous and cumulative downward fluxes calculated in this study likely under-estimated the true magnitude of downward diffusion. Future work could better understand this possibility by installing deeper CO2 sensors (Benavente et al., 2010), simultaneously instrumenting soil profiles and underlying caves (Meyer et al., 2014), or using isotope discrimination techniques to trace carbon movement throughout the subsurface (Breecker et al., 2012; Mattey et al., 2016). Due to limitations in materials and budget, the study was limited to a single site, with only one set of sensors per sinkhole slope position. The lack of replication prevented statistical comparisons between the different soil profiles or with other sinkholes. Nonetheless, the similar range of concentrations and seasonal patterns measured in all three profiles suggests that the sensors were adequately capturing CO2 dynamics within the sinkhole. Measured CO2 concentrations and calculated fluxes also tended to be similar between years, which indicates that sensor drift was not a major issue during the study.
226
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
Controls on CO2 Fluxes in Karst
IMPLICATIONS AND CONCLUSIONS Results of this study support previous work and also offer new insights into CO2 gas movement within karst soils. For example, similar to previous work, we detected increasing soil CO2 concentrations during the growing season, leading to peak upward fluxes in the middle of the growing season, followed by decreased concentrations during the dormant season. These seasonal patterns have been noted in many investigations in both karst and non-karst soils (e.g., Yang et al., 2012). However, our study provided additional unique findings. In particular, the use of multiple sensors allowed us to analyze both the magnitudes and directions of CO2 diffusion in karst soils. Our data set included many periods in which measured CO2 concentrations were lower at 60 cm compared to 20 cm depth. This condition implies the presence of a near-surface zero-flux plane for CO2, similar to the zero-flux planes that can be seen for subsurface hydrologic gradients (e.g., Brutsaert, 2014). It is probable that downward gas diffusion occurred during these periods (assuming non-zero gas diffusivity), which is a process that has not been carefully considered prior to this study. We posited that the underlying cave system acts as a primary CO2 sink in the subsurface, especially during the cool season, when convective venting likely occurs. This process allows relatively rapid exchange between air in the cave and the external atmosphere (Mattey et al., 2016; Krajnc et al., 2017). The relatively low CO2 concentrations that occur during these periods (Meyer et al., 2014; Kukuljan et al., 2021) would encourage downward diffusion through the soil profile. Future work could incorporate observations such as the ones collected in this study into biogeochemical models, and thereby better understand the relative importance of each of these mechanisms. Finally, doline caves like James Cave occur throughout southwest Virginia and the Appalachian karst system. While our study was limited to a single site, it is likely that the findings of this study are generalizable to other karst areas with similar geologic and environmental factors. In particular, our study results suggest that CO2 gas exchange within soil profiles above cave systems is highly dynamic over space and time, influenced by soil properties such as temperature, depth, water content, and organic matter, but also properties of the underlying cave. The magnitude of upward fluxes was determined to have similar seasonal trends between soil profiles, with most of the CO2 efflux occurring during the warm growing season. The toe-slope profile had greater cumulative efflux than the back slope and shoulder, which was likely caused by connected air-filled pore space. Air-filled porosity was also important for downward gas diffusion, which was a much smaller but not negligible flux in the toe slope. Our study also emphasized the importance of quantifying soil hydraulic properties, such as air-filled porosity (e), when
determining subsurface CO2 fluxes. This type of characterization should be used in future work to constrain terrestrial carbon budgets and quantify carbon storage within karst areas. ACKNOWLEDGMENTS Funding for this work was provided the Global Change Center at Virginia Tech and by the Virginia Agricultural Experiment Station and the Hatch Program of the U.S. Department of Agriculture National Institute of Food and Agriculture (1026126). Support was also provided by the George Washington Carver program at Virginia Tech. We acknowledge Wil Orndorff, Daniel Doctor, and Benjamin Schwartz for discussions during the development of this project, and we thank Brandon Lester, Ayush Gyawali, Jesse Radolinski, Jinshi Jian, Caroline Wolcott, and Jonathan Grubb for their help with fieldwork and laboratory analyses. SUPPLEMENTAL MATERIAL Supplemental Material associated with this article can be found online at https://doi.org/10.2113/EEG-D-22-00082. REFERENCES ALI, R. S.; POLL, C.; AND KANDELER, E., 2018, Dynamics of soil respiration and microbial communities: Interactive controls of temperature and substrate quality: Soil Biology and Biochemistry, Vol. 127, pp. 60–70. https://doi.org/10.1016/j.soilbio.2018.09.010 BENAVENTE, J.; VADILLO, I.; CARRASCO, F.; SOLER, A.; LIÑÁN, C.; AND MORAL, F., 2010, Air carbon dioxide contents in the vadose zone of a Mediterranean karst: Vadose Zone Journal, Vol. 9, No. 1, pp. 126–136. BREECKER, D. O.; PAYNE, A. E.; QUADE, J.; BANNER, J. L.; BALL, C. E.; MEYER, K. W.; AND COWAN, B. D., 2012, The sources and sinks of CO2 in caves under mixed woodland and grassland vegetation: Geochimica et Cosmochimica Acta, Vol. 96, pp. 230–246. BRUTSAERT, W., 2014, The daily mean zero-flux plane during soilcontrolled evaporation: A Green’s function approach: Water Resources Research, Vol. 50, No. 12, pp. 9405–9413. https://doi.org/10.1002/ 2014wr016111 CAO, M.; JIANG, Y.; CHEN, Y.; FAN, J.; AND HE, Q., 2020, Variations of soil CO2 concentration and pCO2 in a cave stream on different time scales in subtropical climatic regime: Catena, Vol. 185, p. 104280. CHEN, Q., 2019, Characteristics of soil profile CO2 concentrations in karst areas and their significance for global carbon cycles and climate change: Earth System Dynamics, Vol. 10, No. 3, pp. 525–538. COWAN, B. D.; OSBORNE, M. C.; AND BANNER, J. L., 2013, Temporal variability of cave-air CO2 in central Texas: Journal of Cave and Karst Studies, Vol. 75, No. 1, pp. 38–50. CUEZVA, S.; FERNANDEZ-CORTES, A.; BENAVENTE, D.; SERRANO-ORTIZ, P.; KOWALSKI, A. S.; AND SANCHEZ-MORAL, S., 2011, Short-term CO2(g) exchange between a shallow karstic cavity and the external atmosphere during summer: Role of the surface soil layer: Atmospheric Environment, Vol. 45, No. 7, pp. 1418–1427. https://doi. org/10.1016/j.atmosenv.2010.12.023 DANIEL, J.; POTTER, K.; ALTOM, W.; ALJOE, H.; AND STEVENS, R., 2002, Long-term grazing density impacts on soil compaction: Transactions of the ASAE, Vol. 45, No. 6, p. 1911.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
227
Thompson, McLaughlin, Schreiber, and Stewart EAGLE, S.; ORNDORFF, W.; SCHWARTZ, B.; DOCTOR, D. H.; GERST, J.; AND SCHREIBER, M., 2015, Analysis of hydrologic and geochemical time-series data at James Cave, Virginia: Implications for epikarst influence on recharge in Appalachian karst aquifers. In Feinberg, J. M.; Gao, Y.; and Alexander, E. C., Jr. (Editors), Caves and Karst Across Time: Special Paper 516, Geological Society of America, Boulder, CO, pp. 181–196. FAN, J. AND JONES, S. B., 2014, Soil surface wetting effects on gradient-based estimates of soil carbon dioxide efflux: Vadose Zone Journal, Vol. 13, No. 2, pp. 1–12. FANG, C. AND MONCRIEFF, J. B., 2005, The variation of soil microbial respiration with depth in relation to soil carbon composition: Plant and Soil, Vol. 268, pp. 243–253. https://doi.org/10.1007/s11104004-0278-4 FISSORE, C.; DALZELL, B. J.; BERHE, A.; VOEGTLE, M.; EVANS, M.; AND WU, A., 2017, Influence of topography on soil organic carbon dynamics in a Southern California grassland: Catena, Vol. 149, pp. 140–149. FORD, D. AND WILLIAMS, P. D., 2007, Karst Hydrogeology and Geomorphology: John Wiley & Sons, New York. GROCE-WRIGHT, N. C.; BENTON, J. R.; HAMMOND, N. W.; AND SCHREIBER, M. E., 2022, A decade of cave drip hydrographs shows spatial and temporal variability in epikarst storage and recharge to Appalachian karst systems: Hydrology, Vol. 9, No. 131, pp. 1–15. JAYARATHNE, J. R. R. N.; CHAMINDU DEEPAGODA, T. K. K.; CLOUGH, T. J.; NASVI, M. C. M.; THOMAS, S.; ELBERLING, B.; AND SMITS, K., 2020, Gas-diffusivity based characterization of aggregated agricultural soils: Soil Science Society of America Journal, Vol. 84, No. 2, pp. 387–398. https://doi.org/10.1002/saj2.20033 JIAN, J.; BAILEY, V.; DORHEIM, K.; KONINGS, A. G.; HAO, D.; SHIKLOMANOV, A. N.; SNYDER, A.; STEELE, M.; TERAMOTO, M.; VARGAS, R.; AND BOND-LAMBERTY, B., 2022, Historically inconsistent productivity and respiration fluxes in the global terrestrial carbon cycle: Nature Communications, Vol. 13, No. 1, pp. 1–9. https://doi.org/10.1038/s41467-022-29391-5 JIN, Y. AND JURY, W. A., 1996, Characterizing the dependence of gas diffusion coefficient on soil properties: Soil Science Society of America Journal, Vol. 60, No. 1, pp. 66–71. KOPP, M.; KAYE, J.; SMEGLIN, Y. H.; ADAMS, T.; PRIMKA, E. J.; BRADLEY, B.; SHI, Y.; AND EISSENSTAT, D., 2022, Topography mediates the response of soil CO2 efflux to precipitation over days, seasons, and years: Ecosystems, Vol. 26, pp. 1–19. https://doi.org/10.1007/ s10021-022-00786-1 KRAJNC, B.; FERLAN, M.; AND OGRINC, N., 2017, Soil CO2 sources above a subterranean cave—Pisani rov (Postojna Cave, Slovenia): Journal of Soils and Sediments, Vol. 17, No. 7, pp. 1883–1892. KUKULJAN, L.; GABROVŠEK, F.; COVINGTON, M. D.; AND JOHNSTON, V. E., 2021, CO2 dynamics and heterogeneity in a cave atmosphere: Role of ventilation patterns and airflow pathways: Theoretical and Applied Climatology, Vol. 146, No. 1, pp. 91–109. LAEMMEL, T.; MAIER, M.; SCHACK-KIRCHNER, H.; AND LANG, F., 2017, An in situ method for real-time measurement of gas transport in soil: European Journal of Soil Science, Vol. 68, No. 2, pp. 156– 166. https://doi.org/10.1111/ejss.12412 LAFOND, J. A.; ALLAIRE, S. E.; DUTILLEUL, P.; PELLETIER, B.; LANGE, S. F.; AND CAMBOURIS, A. N., 2011, Spatiotemporal analysis of the relative soil gas diffusion coefficient in two sandy soils: Variability decomposition and correlations between sampling dates at two spatial scales: Soil Science Society of America Journal, Vol. 75, No. 5, pp. 1613–1625. MATTEY, D. P.; ATKINSON, T. C.; BARKER, J. A.; FISHER, R.; LATIN, J. P.; DURRELL, R.; AND AINSWORTH, M., 2016, Carbon dioxide, ground air and carbon cycling in Gibraltar karst: Geochimica et Cosmochimica Acta, Vol. 184, pp. 88–113. https://doi.org/10.1016/j.gca.2016.01.041
228
MCCOURTY, M. A.; GYAWALI, A. J.; AND STEWART, R. D., 2018, Of macropores and tillage: Influence of biomass incorporation on cover crop decomposition and soil respiration: Soil Use and Management, Vol. 34, pp. 101–110. https://doi.org/10.1111/sum.12403 MEYER, K. W.; FENG, W.; BREECKER, D. O.; BANNER, J. L.; AND GUILFOYLE, A., 2014, Interpretation of speleothem calcite d13C variations: Evidence from monitoring soil CO2, drip water, and modern speleothem calcite in central Texas: Geochimica et Cosmochimica Acta, Vol. 142, pp. 281–298. PACIFIC, V. J.; MCGLYNN, B. L.; RIVEROS-IREGUI, D. A.; EPSTEIN, H. E.; AND WELSCH, D. L., 2009, Differential soil respiration responses to changing hydrologic regimes: Water Resources Research, Vol. 45, No. 7, pp. 1–6. PACIFIC, V. J.; MCGLYNN, B. L.; RIVEROS-IREGUI, D. A.; WELSCH, D. L.; AND EPSTEIN, H. E., 2008, Variability in soil respiration across riparian-hillslope transitions: Biogeochemistry, Vol. 91, No. 1, pp. 51–70. PACIFIC, V. J.; MCGLYNN, B. L.; RIVEROS-IREGUI, D. A.; WELSCH, D. L.; AND EPSTEIN, H. E., 2011, Landscape structure, groundwater dynamics, and soil water content influence soil respiration across riparian– hillslope transitions in the Tenderfoot Creek Experimental Forest, Montana: Hydrological Processes, Vol. 25, No. 5, pp. 811–827. PUMPANEN, J.; ILVESNIEMI, H.; AND HARI, P., 2003, A process-based model for predicting soil carbon dioxide efflux and concentration: Soil Science Society of America Journal, Vol. 67, No. 2, pp. 402–413. RIECHELMANN, S.; BREITENBACH, S. F. M.; SCHRÖDER-RITZRAU, A.; MANGINI, A.; AND IMMENHAUSER, A., 2019, Ventilation and cave air PCO2 in the Bunker-Emst Cave System (NW Germany): Implications for speleothem proxy data: Journal of Cave and Karst Studies, Vol. 81, No. 2, pp. 98–112. https://doi.org/10.4311/2018ES0110 SAGGAR, S.; MACKAY, A.; AND HEDLEY, C., 1999, Hill slope effects on the vertical fluxes of photosynthetically fixed 14C in a grazed pasture: Soil Research, Vol. 37, No. 4, pp. 655–666. SANCHEZ-CAÑETE, E.; SERRANO-ORTIZ, P.; KOWALSKI, A.; OYONARTE, C.; AND DOMINGO, F., 2011, Subterranean CO2 ventilation and its role in the net ecosystem carbon balance of a karstic shrubland: Geophysical Research Letters, Vol. 38, No. 9, pp. 1–4. SCHREIBER, M. E.; SCHWARTZ, B. F.; ORNDORFF, W.; DOCTOR, D. H.; EAGLE, S. D.; AND GERST, J. D., 2015, Instrumenting caves to collect hydrologic and geochemical data: Case study from James Cave, Virginia. In Younos, T. and Parece, E. E. (Editors), Advances in Watershed Science and Assessment: Springer, Cham, Switzerland, pp. 205–231. Soil Survey Staff, 2019, U.S. Department of Agriculture Soil Survey: Natural Resources Conservation Service, U.S. Department of Agriculture. Electronic document, available at http://websoilsurvey.sc .egov.usda.gov/ TANG, J.; BALDOCCHI, D. D.; QI, Y.; AND XU, L., 2003, Assessing soil CO2 efflux using continuous measurements of CO2 profiles in soils with small solid-state sensors: Agricultural and Forest Meteorology, Vol. 118, No. 3–4, pp. 207–220. https://doi.org/10.1016/ s0168-1923(03)00112-6 TIAN, Q.; WANG, D.; TANG, Y.; LI, Y.; WANG, M.; LIAO, C.; AND LIU, F., 2019, Topographic controls on the variability of soil respiration in a humid subtropical forest: Biogeochemistry, Vol. 145, No. 1, pp. 177–192. Vaisala, 2017, CO2 Measurement in Incubators—Questions and Answers: Vaisala. Electronic document, available at https://www.vaisala.com/ sites/default/files/documents/CO2-Measurement-in-IncubatorsApplication-note-B210826EN.pdf WEBSTER, K.; CREED, I.; BOURBONNIERE, R.; AND BEALL, F., 2008, Controls on the heterogeneity of soil respiration in a tolerant hardwood forest: Journal of Geophysical Research: Biogeosciences, Vol. 113, pp. 1–15. WEISBROD, N.; DRAGILA, M. I.; NACHSHON, U.; AND PILLERSDORF, M., 2009, Falling through the cracks: The role of fractures in Earthatmosphere gas exchange: Geophysical Research Letters, Vol. 36, No. 2, pp. 1–5.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
Controls on CO2 Fluxes in Karst WHITE, W. B., 2016, Science of caves and karst: A half century of progress. In Feinberg, J. M.; Gao, Y.; and Alexander, E. C., Jr. (Editors), Caves and Karst Across Time: Special Paper 516, Geological Society of America, Boulder, CO, pp. 19–33. YANG, R.; LIU, Z.; ZENG, C.; AND ZHAO, M., 2012, Response of epikarst hydrochemical changes to soil CO2 and weather conditions at
Chenqi, Puding, SW China: Journal of Hydrology, Vol. 468–469, pp. 151–158. https://doi.org/10.1016/j.jhydrol.2012.08.029 ZHAO, R.; LIU, Z.; DONG, L.; ZHANG, Q.; AND LIU, C., 2021, The fates of CO2 generated by H2SO4 and/or HNO3 during the dissolution of carbonate and their influences on the karst-related carbon cycle: Journal of Hydrology, Vol. 597, p. 125746.
Environmental & Engineering Geoscience, Vol. XXIX, No. 3, August 2023, pp. 217–229
229
EDITORIAL OFFICE: Environmental & Engineering Geoscience journal, Kent State University, Kent, OH 44242, U.S.A., ashakoor@kent.edu. CLAIMS: Claims for damaged or not received issues will be honored for 6 months from date of publication. AEG members should contact AEG, 3053 Nationwide Parkway, Brunswick, OH 44212. Phone: 844-331-7867. GSA members who are not members of AEG should contact the GSA Member Service center. All claims must be submitted in writing. POSTMASTER: Send address changes to AEG, 3053 Nationwide Parkway, Brunswick, OH 44212. Phone: 844-331-7867. Include both old and new addresses, with ZIP code. Canada agreement number PM40063731. Return undeliverable Canadian addresses to Station A P.O. Box 54, Windsor, ON N9A 6J5 Email: returnsil@imexpb.com. DISCLAIMER NOTICE: Authors alone are responsible for views expressed in articles. Advertisers and their agencies are solely responsible for the content of all advertisements printed and also assume responsibility for any claims arising therefrom against the publisher. AEG and Environmental & Engineering Geoscience reserve the right to reject any advertising copy. SUBSCRIPTIONS: Member subscriptions: AEG members automatically receive digital access to the journal as part of their AEG membership dues. Members may order print subscriptions for $75 per year. GSA members who are not members of AEG may order for $60 per year on their annual GSA dues statement or by contacting GSA. Nonmember subscriptions are $310 and may be ordered from the subscription department of either organization. A postage differential of $10 may apply to nonmember subscribers outside the United States, Canada, and Pan America. Contact AEG at 844-331-7867; contact GSA Subscription Services, Geological Society of America, P.O. Box 9140, Boulder, CO 80301. Single copies are $75.00 each. Requests for single copies should be sent to AEG, 3053 Nationwide Parkway, Brunswick, OH 44212. © 2023 by the Association of Environmental and Engineering Geologists All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or by any information storage and retrieval system, without permission in writing from AEG.
EDITORS ABDUL SHAKOOR Kent State University Kent, OH 44242 ashakoor@kent.edu
ERIC PETERSON Department of Geography, Geology, and the Environment Illinois State University Normal, IL 61790 309-438-5669 ewpeter@ilstu.edu
KAREN E. SMITH, Editorial Assistant, kesmith6@kent.edu
OOMMEN, THOMAS Board Chair, Michigan Technological University SASOWSKY, IRA D. University of Akron
ASSOCIATE EDITORS ACKERMAN, FRANCES Ramboll Americas Engineering Solutions, Inc. BASTOLA, HRIDAYA Lehigh University BEGLUND, JAMES Montana Bureau of Mines and Geology BRUCKNO, BRIAN Virginia Department of Transportation CLAGUE, JOHN Simon Fraser University, Canada DEE, SETH University of Nevada, Reno FRYAR, ALAN University of Kentucky GARDNER, GEORGE Massachusetts Department of Environmental Protection GHOSH, SUMAN California Department of Conservation
HAUSER, ERNEST Wright State University KEATON, JEFF WSP USA MAY, DAVID USACE-ERDC-CHL POPE, ISAAC Book Review Editor SANTI, PAUL Colorado School of Mines SCHUSTER, BOB SHLEMON, ROY R.J. Shlemon & Associates, Inc. STOCK, GREG National Park Service SWANSON, SUSAN (SUE) Beloit College ULUSAY, RESAT Hacettepe University, Turkey WEST, TERRY Purdue University
Environmental & Engineering Geoscience AUGUST 2023
VOLUME XXIX, NUMBER 3
Special Issue on Karst: Guest Editors – Cory Blackeagle, Russell Harmon, and Robert Denton
Submitting a Manuscript Environmental & Engineering Geoscience (E&EG), is a quarterly journal devoted to the publication of original papers that are of potential interest to hydrogeologists, environmental and engineering geologists, and geological engineers working in site selection, feasibility studies, investigations, design or construction of civil engineering projects or in waste management, groundwater, and related environmental fields. All papers are peer reviewed. The editors invite contributions concerning all aspects of environmental and engineering geology and related disciplines. Recent abstracts can be viewed under “Archive” at the web site, “http://eeg.geoscienceworld.org”. Articles that report on research, case histories and new methods, and book reviews are welcome. Discussion papers, which are critiques of printed articles and are technical in nature, may be published with replies from the original author(s). Discussion papers and replies should be concise. To submit a manuscript go to https://www.editorialmanager.com/EEG/ default.aspx. If you have not used the system before, follow the link at the bottom of the page that says New users should register for an account. Choose your own login and password. Further instructions will be available upon logging into the system. Upon submission, manuscripts must meet, exactly, the criteria specified in the revised Style Guide found at https://www.aegweb.org/e-eg-supplements. Manuscripts of fewer than 10 pages may be published as Technical Notes. The new optional feature of Open Access is available upon request for $750 per article. For further information, you may contact Dr. Abdul Shakoor at the editorial office.
Cover photo To understand carbon dioxide (CO2) fluxes in karst soils, three soil profiles were instrumented with CO2, temperature, and water content sensors near the entrance to James Cave, Virginia. Photo credit: Taryn Thompson. See article on page 217.
Volume XXIX, Number 3, August 2023
THIS PUBLICATION IS PRINTED ON ACID-FREE PAPER
ADVISORY BOARD WATTS, CHESTER “SKIP” F. Radford University HASAN, SYED University of Missouri, Kansas City NANDI, ARPITA East Tennessee State University
ENVIRONMENTAL & ENGINEERING GEOSCIENCE
Environmental & Engineering Geoscience (ISSN 1078-7275) is published quarterly by the Association of Environmental & Engineering Geologists (AEG) and the Geological Society of America (GSA). Periodicals postage paid at AEG, 3053 Nationwide Parkway, Brunswick, OH 44212 and additional mailing offices.
THE JOINT PUBLICATION OF THE ASSOCIATION OF ENVIRONMENTAL AND ENGINEERING GEOLOGISTS AND THE GEOLOGICAL SOCIETY OF AMERICA SERVING PROFESSIONALS IN ENGINEERING GEOLOGY, ENVIRONMENTAL GEOLOGY, AND HYDROGEOLOGY