Managing Marshes with Drawdowns Can Reduce Elevation of Coastal Marshes

doi:10.21203/rs.3.rs-3559501/v1

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Managing Marshes with Drawdowns Can Reduce Elevation of Coastal Marshes

https://doi.org/10.21203/rs.3.rs-3559501/v1

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Coastal wetlands can offset some subsidence and effects of sea level rise via vertical accretion. We used ¹³⁷Cs dating to compare 55 years of accretion between an impounded marsh and an un-impounded marsh. The un-impounded marsh had five times more accretion. In the un-impounded marsh, accretion was related to organic accumulation whereas mineral accumulation was related to bulk density. In the impounded marsh, accretion was related to mineral accumulation and bulk density. We attributed those differences to natural and managed prolonged drainage events since ~2005 in the impounded marsh that we studied.

We compared our estimates to earlier 33-year estimates using ¹³⁷Cs from the same un-impounded marsh. Those estimates were slower for accretion, mineral accumulation, and organic accumulation. We estimated that accretion would have had to accelerate 68%, organic accumulation accelerate 11%, and mineral accumulation accelerate 7-fold after 1998 for us to observe the rates that we estimated. We attributed those differences to increases in flooding by tides and/or tropical storms since 2000.

We compared our estimates to earlier 33-year estimates using ¹³⁷Cs from two nearby impounded marshes. Impoundment effects varied widely, which precludes broad statements about effects of impoundment on accretion. None-the-less, evidence is accumulating that moist-soil management on organic soils in the coastal zone compromises accretion processes and reduces elevation via soil organic matter oxidation and compaction. New research is suggested on highly organic soils to identify water management strategies to reduce decomposition and to identify plant species and fire and water management strategies that foster accretion.

coastal marsh

vertical accretion

wildlife management

impoundment

Waterfowl conservation began in the early 20th century as populations declined due to market hunting and a loss of wetlands due to agriculture and flood control (Trefethen 1961). A variety of federal laws, including Migratory Bird Hunting Act in 1934, Pittman Robertson Wildlife Restoration Act of 1937, and the North American Wetland Conservation Act in 1989 as well as conservation efforts by state, federal, and non-governmental organizations supported the establishment of regulations, habitat protection, and habitat management techniques (Trefethen 1961), The North American Waterfowl Management Plan, originally developed in 1986, established a goal of maintaining waterfowl populations at levels observed during the 1970’s (U.S. Fish and Wildlife Service and Canadian Wildlife Service 1986).

The original North American Waterfowl Management Plan recognized the widespread degradation of wetlands and emphasized the importance of protection, restoration, and management of wetlands to reach waterfowl population goals. Moist-soil management is a common wetland management strategy that developed prior to the North American Waterfowl Management plan and is used broadly across inland and coastal systems. The technique generally uses seasonal drawdowns and reflooding to promote annual plants and invertebrates that develop on recently drained mudflats and thereby improve habitat for waterfowl and other waterbirds (Fredrickson and Taylor 1982). The drawdowns stimulate seed germination of numerous species, including mudflat annuals that have high seed production and are an important food source for many waterfowl species. The drawdowns can be quickly followed by shallow reflooding to irrigate more flood-tolerant plants and stress less flood-tolerant plants; drawdowns can persist long into the growing season.

In coastal wetlands, management using levees and water-control structures to hold and exclude water was developed by the 1940s, and by the 1960s, seasonal drawdowns were used in tidal freshwater marshes (e.g., Chabreck 1960; Morgan et al. 1975). However, when used in organic soils, which are common in many coastal marshes ranging from macrotidal temperate (e.g., Chmura and Hung 2004) to microtidal tropical (e.g., McKee et al. 2007), drainage can oxidize organic matter and reduce elevation. Drexler et al. (2013) collected a pair of cores from six study sites in coastal South Carolina. Two sites were open to the tides; two sites were managed with drawdowns that persisted most of the growing season and routinely managed with disking, burning and herbicides; two sites were managed with brief drawdowns early in the growing season. Drexler et al. (2013) measured four times more vertical accretion in the two tidal sites than in the four moist-soil units. The long-term use of seasonal drawdowns not only slows vertical accretion but can reverse many decades of accretion and lead to the loss of a meter or more of elevation across hundreds to thousands of hectares (e.g., Suisun Principal Agencies 2013). Thus, a better understanding of the effects of marsh management on accretion and decomposition processes is needed to sustain long-term marsh health, particularly in face of increasing sea level rise.

The actual processes dictating marsh elevation change are complex and include flooding and draining, shallow and deep subsidence, sediment deposition, and below ground plant production and decomposition processes (Lynch et al. 2015). Local subsidence and global sea-level rise have combined to reduce the elevation of most coastal marshes, but such submergence can be offset for thousands of years by a process called vertical accretion. Vertical accretion can happen biogenically through the accumulation of roots and rhizomes (McCaffrey and Thomson 1980), and/or through the accumulation of mineral sediments delivered by riverine and marine processes (Reed 1989; Roberts et al. 1989; Rosen and Xu 2011; Williams 2011). Nyman et al. (2006) found that below ground production and decomposition drive accretionary processes in Louisiana coastal marshes. Whereas biologists began identifying plant (McAtee 1914) and animal (Wetmore 1925) foods of water birds in the early 1900s, biologists have not yet determined if some plant species promote accretion more than others. Sasser et al. (2018) found that dead and live root biomass represent up to 22% of the total mass of soil solids in Louisiana coastal marshes. Thus, factors affecting below ground production, such as salinity and flood duration (Snedden et al. 2015), and differences between annual and perennial species in root biomass (Grime 1977; Baldwin 2013), may affect accretion. While flooding indirectly affects accretion through effects on below ground plant productivity, flooding can also cause accretion when water deposits mineral matter or organic debris on the existing marsh surface, or when flooding induces marsh vegetation to produce adventitious water roots over the existing marsh surface (Nyman et al. 2006). In addition to those surface processes, surface elevation also is affected by processes that create expansion (i.e., dilation) or a rise in the soil surface and both shallow and deep subsidence (Lynch et al. 2015). Dilation/expansion occurs when older soil expands upward, apparently in response to groundwater saturation (Nuttle et al. 1990; Chambers et al. 2019). Shallow processes such as autocompaction/shallow subsidence, occur when the weight of newer soil compresses underlying older soil (Kaye and Barghoorn 1964). Shallow subsidence can also occur when deposited organic matter decomposes slowly under anaerobic soil conditions, or more quickly under aerobic conditions, both causing surface elevation loss (Lynch et al. 2015). Deep subsidence is a result of Holocene sediment degassing, compressing, and settling over the older, more stable Pleistocene sediments (Kaye and Barghoorn 1964; Lynch et al. 2015; Allen 1999). Managers cannot alter deep subsidence and some shallow subsidence processes, but they can affect surface processes and some shallow processes that drive marsh elevation.

The management actions that may affect marsh accretion are numerous and generally poorly quantified. Moist-soil management favors annual plants over perennial plants, which differ in root traits in ways that may affect root production and decomposition. The roots of annual plants have low density, short-lived root networks whereas perennial plants generally have higher density, and longer-lived root networks (Roumet et al. 2006). Additionally, the nitrogen concentration is greater in annual plant roots (Roumet et al. 2006), which may increase soil organic decomposition rates via carbon quality and nutrient availability (Sihi et al. 2016). Marsh management also may affect accretion because levees and water-control structures reduce sediment deposition on natural levees during normal tides (Reed 1992). Drawdowns on peat soils initially cause elevation loss through the loss of water in soil pores and compression of the underlying soils (Chambers et al. 2019). The oxygenated soil then begins to decompose faster (Nyman and DeLaune 1991). In a tropical, non-coastal peatland soil 35–100% of elevation loss following drawdown is due to oxidation of organic matter (Couwenberg et al. (2010) and the loss can be rapid in coastal marshes and swamps (Chambers et al. 2014). Salinity can increase or decrease belowground production ranging from short-term increases in decomposition to persistent reductions in decomposition rates (Neubauer et al. 2013) such that salt-water intrusion reduces elevation (Charles et al. 2019). Lane et al. (2016) found that herbicides reduced marsh elevation by 4.24 cm, 1.56 cm, and 1.48 cm in fresh, brackish, and saline marshes, respectively 1.5 years after treatment.

In this study, we evaluate the long-term effects of management, including drawdowns and herbicides, on marsh accretion between a managed and unmanaged impoundment. We hypothesized that the impounded marsh would have slower accretion than the un-impounded marsh.

Site Description

Rockefeller State Wildlife Refuge is located in the Chenier Plain of southwest Louisiana along the Gulf of Mexico coast (Fig. 1). Recreational fishing is allowed in limited areas but not hunting. When the state of Louisiana formally acquired the refuge in 1920 it consisted of approximately 34,803 ha but by 1980 it consisted of about 31,032 ha mainly due to shoreline erosion along the Gulf Coast and some interior marsh loss (Wicker et al. 1983). Bordered by the Gulf of Mexico to the south, the refuge runs parallel with 43 km of gulf shoreline. To its north lies 70 km of coastal marsh. The refuge occupies an important pathway for drainage for approximately 292,332 ha of coastal marshes that receive water from the Mermentau River. In the 1930s, the Intracoastal Waterway was dredged east to west through coastal marshes 40 km north of Rockefeller Refuge, which increased freshwater drainage east and west but interrupted freshwater drainage south towards Rockefeller Refuge (Wicker et al. 1983). Throughout the 1900s, flood protection levees were used to protect agricultural areas north of the coastal wetlands from hurricane storm surge further reduced freshwater flow to Rockefeller Refuge. An east-west state highway constructed further restricted freshwater inflow from the north. In the 1950s, canals dredged within the refuge to mine oil and gas altered natural hydrologic processes at the local scale refuge. To minimize impacts of altered hydrology, refuge managers built levees, or used spoil levees created by canal construction for oil and gas development, and installed water-control structures to manage water levels (Wicker et al. 1983). Most of the refuge currently consists of managed impoundments, in which public access is not allowed. Public access is allowed on approximately 12,000 ha that are not impounded and retain natural tidal connections to the Gulf of Mexico (Fig. 2).

Unit 2 is a managed impounded unit encompassing ~ 345 ha (~ 850 acres) that was built in 1963 with levees on all sides and a water-control structure and pump to regulate hydroperiod (Fig. 2). Since the 1940s, the vegetation has been classified as intermediate (often a species rich Spartina patens marsh) or brackish (generally dominated by S. patens) (https://lacoast.gov/crms_viewer/Map/CRMSViewer). Since construction, it has been surrounded on its west (privately owned), north, and east sides by canals that border other impounded units that also are leveed and equipped with water control structures for gravity drainage or pumps for forced drainage. Unit 2 is bordered on its south by the Price Lake Unit. The unit to the south (Price Lake Unit) has levees on it north (Unit 2, our study site), east and west, but is bounded by the Gulf of Mexico to the south. Although not impounded, water levels in the Price Lake Unit are managed via gravity drainage (Melancon 2000).

From construction until ~ 2005, Unit 2 was managed for nesting American alligators (Alligator mississippiensis) only with gravity drainage, which allowed for relatively brief drainage events, known as drawdowns, that generally kept soils saturated to within a few centimeters of the soil surface. Since ~ 2005 however, Unit 2 has been managed for migrating shorebirds, wading birds, and waterfowl with frequent and persistent drawdowns. The change in management techniques coincided with a return to mostly droughty conditions from 2005 through 2012, which were less prevalent from the 1970s to 2005, but were more persistent from the 2006 until the now (https://www.drought.gov/historical-information?state=louisiana&dataset=1&selectedDateUSDM=20111108&selectedDateSpi=19850701&dateRangeSpi=1950-2020). Unit 2 was drained continually from October 2005 through May 2009 and most of the time through July 2012. Since 2007, it also has been managed to favor surface inputs with saline Gulf water and with herbicides.

The un-impounded marsh consists of ~ 10,800 ha (26,700 acres) that spans almost 25 km (> 40 miles) of shoreline on the Gulf of Mexico. It extends ~ 5 km inland on its western and eastern ends but narrows to only few km in between (Fig. 2). It lacks levees and water-control structures to manage water levels. It lacks artificial tidal connections to the Gulf of Mexico and has only a few natural connections, which is the Chenier Plain (McBride et al. 2007). Lightning occasionally starts wildfires in the un-impounded marsh, but prescribed fire is rarely used there. From the 1940s through the 1960s, its vegetation was classified mostly as brackish but with saline (dominated by Spartina alterniflora) near the shoreline and intermediate on the north (https://lacoast.gov/crms_viewer/Map/CRMSViewer). It has lacked intermediate marsh since the 1960s and has been primarily a saline marsh since 2007 (https://lacoast.gov/crms_viewer/Map/CRMSViewer).

We used ¹³⁷Cs dating to study soil formation: vertical accretion, mineral accumulation, and organic accumulation. Soil cores were collected and processed following methods similar to Nyman et al. (2006). Within each unit, ArcMap Software (ArcMap Version 10.7.1, Environmental Systems Research Institute, Inc., Redlands, CA, USA) was used to create a numbered grid overlay and then SAS Software was used to randomly select grids for coring sites. Within a randomly selected grid, collection sites had to meet the following criteria: support emergent vegetation, be >25 meters away from levees, and be >10 m from open water or ponded areas lacking emergent vegetation.

Soil cores were collected from 10 sites in Unit 2 (Fig. 3) and from 10 sites in the un-impounded marsh (Fig. 4) between October 14, 2018, and September 5, 2019. To collect cores, a 50-cm long or 70-cm long, 15.2-cm diameter PVC pipe was pushed into the marsh surface until refusal or until compaction. To estimate compaction, we recorded in situ measurements of the distance from the top of the PVC pipe to the soil surface inside and outside the pipe. The top of the pipe was then capped, and hand dug out of the ground. The bottom end of the pipe was then capped, stored horizontally, and transported back to the lab for processing. Horizontal transportation of the cores was intended to reduce compaction, which could skew bulk density and vertical accretion estimates but would not alter mass accumulation estimates. The cores were stored in refrigeration until processed in the laboratory. In the lab, core interior space was measured again for compaction calculations as expansion or compaction can occur after the core is transported and stored. The cores were extracted out of the PVC pipe and cut into 3-cm horizontal sections that were placed on a pre-weighed labeled sheet of aluminum foil.

Bulk density (g/cm³) was determined by oven drying each core section at 60° C to a constant weight (≤ 3% change over a 24-hour period) and using the following equation:

soil bulk density = dry soil mass (g) / soil volume (cm³)

Soil volume per soil core increment was calculated, which was 3-cm in height and with a 7.62-cm radius, using the formula 𝑉 = 𝜋𝑟²ℎ. This allowed for bulk density profiles with depth to be constructed because dry weight and soil volume were known variables. Core sections were then homogenized for further analysis using a Wiley Mini-Mill with a number 20 mesh delivery tube.

The 1963 surface was identified in cores with ¹³⁷Cs dating (DeLaune et al. 1978), which allowed us to estimate vertical accretion (cm yr^-1), mineral accumulation (g m^-2 yr^-1), and organic accumulation (g m^-2 yr^-1). We used ¹³⁷Cs dating because (i) the age of the ¹³⁷Cs maxima was similar to the age of Rockefeller Wildlife Refuge impoundments, which were built in the late 1950’s, and (ii) the ¹³⁷Cs maxima was expected to be deep enough to encompass soils old enough to reflect much of the autocompaction and decomposition that are not reflected in feldspar plots. An average vertical accretion rate was estimated from 1963 to when cores were collected, 2018 or 2019, even though vertical accretion rates may have varied within that time. Surface Elevation Tables (SETs) would have allowed detection of changing accretion over time but would have been blind to all differences in accretion between the impounded and un-impounded marsh prior to installing SETs.

¹³⁷Cs analyses were subcontracted to five laboratories for gamma-ray spectroscopy: Tulane University, Indiana University, Texas A&M University at Galveston, and at two laboratories at Louisiana State University. All laboratories used high purity germanium detectors with three different container geometries: planar, well, and Marinelli beaker. Well detectors are the most efficient in terms of counts/gram, but they accommodate the smallest soil sample and thus detect the fewest counts/hour. Planar detectors are the least efficient and thus detect the fewest counts/gram but accommodate more soil than well detectors and so detect more counts/hour. Marinelli beakers accommodate the entire soil core increment and thus detect the most counts/hour by far; they detect fewer counts/gram than well detectors but more than planar detectors. Count times ranged from 3 to 48 hours because detector and geometric configuration changed the counting efficiency and therefore changed the count times required to detect a ¹³⁷Cs peak. Multiple detectors and multiple geometric configurations were used for this study because identifying the 1963 soil surface required identifying only the cesium peak within each core and not the absolute cesium content. A core was deemed successfully ¹³⁷Cs dated if (i) it had a single clearly identifiable peak, and (ii) the peak was not spread across more than two adjacent soil core sections (Drexler et al. 2013).

We also attempted to use ²¹⁰Pb dating on five cores simultaneously when counting ¹³⁷Cs. Only five cores were dated for Pb²¹⁰ activity with two cores from the impounded marsh and three from the un-impounded marsh. Core sections for Pb²¹⁰ were counted using 5-15 g of sample packed in a 60 X 15 mm polystyrene petri dish on a planar detector for 24 hours. To increase the likelihood of reaching ²¹⁰Pb background levels as well as not detecting a false ¹³⁷Cs peak, three of the five cores were 70 cm long to reach deeper depths, one in the impounded unit and two in the un-impounded unit.

Organic and mineral contributions to the soil were determined by ashing a ~5mg sub sample of each core section in a muffle furnace at 400° C for 16 hours. This method of lower temperature ignition does not remove structural water in the clay lattice but does require longer ignition time (Ball 1964; Hesse 1971). Mineral contributions were calculated by weighing the pre-combusted sample and then determining the post-combusted sample. Organic density (g/cm³) and mineral density (g/cm³) of each increment were estimated by multiplying bulk density by percent organic and mineral matter, respectively. Mineral accumulation and organic accumulation since 1963 were calculated using the following equations (Nyman et al. 2006):

Accumulation rates were estimated by dividing accumulation estimates by the number of years between 1963 and the year the core was collected. Average bulk density was estimated by dividing the bulk accumulation by the depth of the ¹³⁷Cs maxima.

Estimates of central tendency and dispersion were reported as mean and standard error. We used correlation coefficients to test for associations among soil parameters. We compared soil parameters between Unit 2 and un-impounded marsh as a Completely Randomized Design with replication. Average annual vertical accretion rate, and organic and mineral accumulation rates since 1963 between the impounded and un-impounded marsh units were compared using SAS Proc Mixed. We used Multiple Regression with backward elimination to determine the relative contribution of mineral and organic matter to vertical accretion and bulk density. Bulk density between the units was compared using SAS Proc Mixed. Mineral and organic contributions to accretion was analyzed using SAS Proc Reg. SAS software was used to compute all statistical analyses. All statistical tests of significance used a 0.05 alpha level.

Vertical accretion in the 20 cores ranged from 0.08 cm yr^-1 to 0.78 cm yr^-1 but the slowest rate in the un-impounded marsh was faster than the fastest rate in the impounded marsh (Table 1). Vertical accretion, mineral accumulation and organic accumulation all were slower in the impounded marsh than in the un-impounded marsh whereas bulk density was higher in the impounded marsh (P < 0.0001 for each of parameters). Vertical accretion averaged 0.11 cm yr^-1 (0.01) in the impounded marsh but 0.56 cm yr^-1 (0.03) in the un-impounded marsh. Mineral accumulation averaged 468.82 g m^-2 yr^-1 (53.6) in the impounded marsh but 1,050.2 g m^-2 yr^-1 (113.0) in the un-impounded marsh. Organic accumulation in the impounded marsh averaged 107.9 g m^-2 yr^-1 (14.2) but averaged 322.5 g m^-2 yr^-1 (16.9) in the un-impounded marsh. Soil bulk density above the 1963 horizon also varied significantly between the impounded and un-impounded marsh units (P < .0001). The un-impounded marsh unit had lower soil bulk density (P), 0.24 g cm^-3 (0.02), than the impounded marsh 0.52 g cm^-3 (0.02).

On average, one centimeter of soil in the un-impounded marsh contained 1,808.0 g m^-2 (157.1) of mineral sediments, but one centimeter soil in the impounded marsh contained 4,205.2 g m^-2 (263.7) mineral sediments. A centimeter of soil in both units contained much less organic matter than mineral matter. Soil in the un-impounded marsh contained 588 g m^-2 (22.8) grams of organic matter but soil in the impounded marsh contained 987.7 g m^-2 (109.1).

Vertical accretion was correlated to mineral accumulation (r = 0.82160, P < 0.0001) and organic accumulation (r = 0.95185, P < 0.0001). Likewise, mineral accumulation and organic accumulation were correlated to each other (r = 0.90152, p < 0.0001). Bulk density was negatively correlated to mineral accumulation (r = -0.49856, P = 0.0253) and vertical accretion (-0.84242, P = < 0.0001), but positively to organic accumulation (r = 0.70244, P = 0.0006.

We used Multiple Regression with backward elimination to determine if mineral or organic matter limited accretion and bulk density in these marshes. We analyzed the impounded and un-impounded marshes separately because accretion, mineral accumulation, organic accumulation, and bulk density differed between these marshes.

Within the impounded unit, accretion was unrelated to organic accumulation (P = 0.7879) and was removed from the model and accretion was related only to mineral accumulation (P = 0.0029). The significant relationship between accretion and mineral accumulation does not prove that mineral matter caused accretion but the lack of a relationship between accretion and organic accumulation precludes organic accumulation from contributing significantly to accretion in the impounded unit.

The reverse occurred in the un-impounded marsh where accretion was unrelated to mineral accumulation (P = 0.3321) and was removed from the model and accretion was related only to organic accumulation (P = 0.0029). The significant relationship between accretion and organic accumulation does not prove that organic matter caused accretion but the lack of a relationship between accretion and mineral accumulation precludes mineral accumulation from contributing significantly to accretion in the un-managed unit.

One of the five cores to be dated with Pb²¹⁰ was not used because there was no pattern, which suggested bioturbation or storm mixing. That core was from the un-impounded marsh. The two cores from the impounded marsh, H and I, had vertical accretion rates of 0.20 cm yr^-1 and 0.26 cm yr^-1, respectively. Neither of the other cores from the un-impounded marsh were deep enough to reach background Pb²¹⁰ levels even though we collected cores almost 70-cm long. Thus, excess Pb²¹⁰ activity could not definitively be calculated. Assuming that the background activity is lower than activity measured in the core and that background levels decline linearly, core G and J vertical accretion rates were 0.47 cm yr^-1 and 0.59 cm yr^-1, respectively.

The results of this study indicate that long-term marsh management practices in a coastal marsh with organic soils can reduce vertical accretion relative to an un-impounded marsh. Our results build upon previous studies that used different techniques, lower sampling intensity, and/or were conducted in different regions (e.g., Cahoon 1994; Bryant and Chabreck 1998; Drexler et al. 2013). We concluded that mineral accumulation was related to bulk density in both marshes, but the two marshes differed in accretion processes. In the impounded marsh, accretion was related only to mineral accumulation whereas in the un-impounded marsh, accretion was related only to organic accumulation. These observations agree with previous researchers who concluded that mineral sediments generally control bulk density (Hatton et al. 1983), that vertical accretion is related to mineral sediments in some marshes (e.g., Temmerman et al. 2003) but that vertical accretion is related to organic matter in other marshes (e.g., Nyman et al. 2006). The accretion rates that we observed in un-impounded marshes were slower than observed in southeast Louisiana (Nyman et al. 1991; Nyman et al. 2006) but faster than rates observed westward in coastal Texas (Moon et al. 2002) probably because subsidence rates decline from east to west in this region (Morton et al. 2006). Accretion increases as subsidence increases partly because flooding increases the opportunity for mineral sediments to be carried onto the marsh, partly because flooding slows decomposition on the marsh, and partly because flooding increases production of water roots at the marsh surface (Nyman et al. 2006). If flooding increases too quickly, then flooding induces a system in which flooding slows root production, which slows accretion, which slows root production, etc. (Nyman et al. 1994).

The most likely reasons that the impounded marsh had much slower accretion than the un-impounded marsh is that impounded marsh (i) depended upon mineral accumulation for vertical accretion, (ii) received only about half the mineral sediments as the un-impounded marsh, and (iii) created soil with twice as much bulk density as the un-impounded marsh. Thus, the impounded marsh had a smaller amount of mineral matter that was compacted into a smaller volume. This caused the bulk density to be over twice as high in impounded marsh even though mineral accumulation in the impounded marsh accumulated less than half as much mineral sediments as the un-impounded marsh. Thus, the impounded marsh required 4,205 g m^-2 of mineral matter to create one centimeter of soil.

The un-impounded marsh was able to accrete five times more elevation from 1963 to 2019 because it depended on organic matter for new soil to create new elevation, and because a centimeter of soil there required only 588 g m^-2 yr^-1 of organic matter. Mineral sedimentation in the un-impounded marsh was directly related to bulk density not to accretion. However, even when mineral matter is not directly related to accretion, it still may contribute indirectly to organic matter accumulation by providing iron, which can help to buffer marsh vegetation from sulfide toxicity (Nyman et al. 1991).

The most likely reason that the managed marsh depended on mineral accumulation for accretion rather than on organic accumulation for the slower vertical accretion in the impounded unit probably is the frequent drawdowns in the managed unit, but saltwater and herbicide treatments to control robust emergent vegetation there may have also increased decomposition of organic matter (Charles et al. 2014; Lane et al 2016). Accelerated soil organic matter decomposition can increase soil compaction and bulk densities when the water level is drawn down and the soil pore spaces dry out and shrink due to oxidation of the organic matter (Chambers et al. 2019). Thus, the highly organic soils of this marsh probably were lost via oxidation when the water level in the unit was drawn down most years during the growing season. Oxidation enables the soil to switch from an anaerobic environment to an aerobic environment, thus speeding up the microbial consumption of organic matter. Chambers et al. (2019), using data from Chambers et al. (2014), found that following drawdown of an unvegetated south Florida organic peat soil, all soil carbon in the upper 2 cm of soil could be consumed in 67 days. Similarly, our analyses of carbon emissions to carbon content data presented in Nyman (1991; Fig. 5) indicate that it would take about 2 months to consume the carbon of 1 cm soil in fresh marsh, one year in brackish marsh, and about 8 months in saline marsh. The impounded marsh was rarely, if ever, drawn down for a year, and accretion, albeit limited, did occur in the marsh. Those mineral sediments must have been deposited before 1963, which suggests that they resulted from storm tides rather than normal wind/tidal flooding. However, vertical accretion in the impounded marsh was near half that estimated global sea-level rise, which is estimated globally at 2.4 mm yr^-1 (Peltier and Tushingham 1989) and at 2.1 mm yr^-1 more recently for south Florida (Khan et al. 2022).

The oxidation of organic material during drawdowns probably was also responsible for soil bulk densities being about twice as much in the impounded unit as in the un-impounded marsh. Despite its lower bulk density, the un-impounded marsh had more mineral contributions than the impounded unit. It is possible that this mineral matter was introduced through strong winter storms (Baumann et al. 1984; Reed 1989) and hurricane passages (Nyman et al. 1995) due to the numerous tidal connections of bayous and creeks, of which the impounded marsh is disconnected.

Our estimates can be used to estimate the amounts of mineral and organic matter accumulation needed to offset any amount of local subsidence and global sea level rise. On average, one centimeter of soil in the un-impounded marsh contained 1,808 g m^-2 (SD = 497) of mineral sediments but one centimeter of soil in the impounded marsh contained 4,205 g m^-2 (SD = 834) of mineral sediments. One centimeter of soil in both units contained much less mass of organic matter than mineral matter. One centimeter of soil in the un-impounded marsh contained 588 g m^-2 (SD = 72) grams of organic matter but one centimeter of soil in the impounded marsh contained 988 g m^-2 (SD = 335) of organic matter.

The results of this study are comparable to Cahoon (1994) who also measured vertical accretion rates in drawdown marshes and un-impounded tidal marsh. Cahoon (1994) used feldspar marker horizons to compare vertical accretion between streamside marsh in the un-impounded marsh at Rockefeller Refuge and a different impounded marsh than we studied: Unit 4. He observed that the un-impounded marsh had significantly higher vertical accretion and organic matter accumulation when compared to Unit 4 during a drawdown. On the east coast, naturally flooded tidal units in South Carolina had four times greater vertical accretion rates, as well as three times less soil bulk density when compared to managed reference marsh (Drexler et al. 2013). Boyd and Sommerfield (2016), noted similar results for their east coast marshes; they found that two impounded marshes had lower rates of vertical accretion when compared to naturally tidal marsh, as well as all marshes having significant correlation between accretion and organic matter accumulation. In contrast, marsh management designed to reverse artificial drainage has been shown to increase carbon storage and vertical accretion by managing for emergent wetland species such as Typha spp. and Phragmites spp., which have dense roots and rhizomes that contribute to organic accretion (Miller et al. 2008; Drake et al. 2015: Table 4).

We did not attempt to corroborate estimates of vertical accretion estimated from peak ¹³⁷Cs concentrations with vertical accretion estimated form Pb²¹⁰ profiles as Drexler et al. (2018) suggested. This is because those estimates could agree only if vertical accretion remained the same over the 55/56 years reflected by ¹³⁷Cs dating and the ~ 80 years reflected by ²¹⁰Pb dating, which is unlikely given that global sea level rise has been accelerating during the last century as has marsh vertical accretion (Kirwan and Megonigal 2013). Even if vertical accretion remained static over time, autocompaction causes accretion estimates from shorter time periods to exceed those from longer time periods (Lynch et al. 2015; Stagg et al. 2016). Thus, such agreement between ¹³⁷Cs dating and ²¹⁰Pb dating will only be possible for cores collected ~ 2040, i.e., ~ 80 years after 1963, but as Drexler et al. (2018), concluded the 1963 peak becomes less detectable because ¹³⁷Cs has a relatively short half-life of 30 years. Such obsolescence cannot be avoided but it may be possible to delay it by collecting more cores, counting larger sample volumes, and using longer count times.

Foret (1997, 2001) also used ¹³⁷Cs dating to study soil formation in the un- impounded marsh (Table 2). His core locations and ours focused on the eastern end of the un-impounded marsh and were similar distances from the Gulf of Mexico (Fig. 5). Comparisons between different time periods can be misleading because of autocompaction. Newer, surface soils compact older, underlying soils (Cahoon et al. 1995; Stagg et al. 2016) such that estimates that include less of the deeper, more compacted soils can appear to have faster accretion and lower bulk density than estimates based on more of the deeper, more compacted soils even when accretion and density were initially the same (Fig. 6). Autocompaction does not affect mineral accumulation rates and organic accumulation rates, which are mass based rather than volume based.

If autocompaction alone was affecting these soils, then our longer time period accretion estimates should have been slower than the shorter time period estimates by Foret (1997, 2001). However, our estimates, based on ~ 32 cm of soil overlying the 1963 marsh surface in 2018/2019, were 26% faster than the older estimates by (Foret 1997; 2001), based on ~ 16 cm of soil overlying 1963 marsh surface in 1996/1998 (Table 2). We estimated that vertical accretion would have had to increase 68% after Foret (1997, 2001) collected his cores (from 0.46 cm yr^− 1 during 1963–1996 to 0.77 cm yr^− 1 during 1996–2019) for us to estimate that vertical accretion averaged 0.58 cm yr^− 1 from 1963 through 2019. Elevation data collected using Surface Elevation Tables since 2007 in the un-impounded marsh suggest that elevation gain accelerated after 2015 (Fig. 7). It is possible that the increase in accretion was driven by sea level in the Gulf of Mexico, which is rising and accelerating, especially in the last 10 years, although with significant regional variation (Ezer 2022). Increasing sea levels would increase flooding, which can slow soil organic matter decomposition and stimulate production of new roots just above the marsh surface (Nyman et al. 2006).

We estimated that mineral accumulation would have had to increase 7.3-fold (from 306 g m^− 2 yr^− 1 to 2,234 g m^− 2 yr^− 1) for us to estimate that mineral accumulation averaged 1,050 g m^− 2 yr^− 1 from 1063 through 2019. The estimated acceleration in organic accumulation was much less; requiring only an increase 11% (from 292 g m^− 2 yr^− 1 to 323 g m^− 2 yr^− 1) for us to estimate that organic accumulation averaged 292 g m^− 2 yr^− 1 from 1963 through 2019. Despite the 7.3-fold increase in mineral accumulation since ~ 1998, vertical accretion from 1963 through 2019 depended on organic accumulation rather than mineral accumulation. The increased mineral accumulation contributed significantly to soil bulk density rather than vertical accretion. Bulk density averaged 0.25 cm^− 3 (std = 0.06) in soils deposited between 1963 and 2018/2019 but averaged 0.14 cm^− 3 (std = 0.01) in soils deposited between 1963 and 1996 (Foret 1998). Foret (2001) did not report average bulk density for soil deposited between 1963 and 1998 but bulk density profiles of those three cores did not exceed 0.2 g cm^− 3.

The increase in mineral accumulation coincided with increasing tropical storm frequency. We searched a list of all named hurricanes that made landfall in the United States (from https://www-aoml-noaa-gov.libezp.lib.lsu.edu/hrd/hurdat/All_U.S._Hurricanes.html) for named hurricanes that made landfall in Louisiana or Texas. We included Texas because most hurricanes that make landfall in Texas cause storm tides in Louisiana because hurricane winds rotate counterclockwise. Those data indicated that named hurricanes increased 25% from 1963–1998 (0.55 hurricanes/year) to 1998–2018 (0.68 hurricanes/year). That 25% increase was accompanied by a decline in strength storminess although the average category declined from a category from 2.2 (STD = to 1.9) to a category 1.8 (STD = 1.2) during that time. None of the 20 hurricanes from 1964 through 1998 made landfall in Louisiana and Texas, but three of the 13 hurricanes from 1998 through 2017 made landfall in both states.

Foret (1997; 2001) also used ¹³⁷Cs dating to study soil formation in two impounded marshes at Rockefeller Refuge (Table 2). Unit 6 encompasses ~ 7,100 ha (17,700 acres); Unit 15 encompasses ~ 300 ha (730 acres). Combining estimates from three impounded marshes indicates that the effects of impoundment vary widely among impoundments. Vertical accretion varied from 0.11 cm yr^− 1 that we observed in Unit 2 from 1963–2019, to 0.50 cm yr^− 1 from 1963–1996 Foret (1997) observed in Unit 15. Differences among estimates could result from differences in time periods between the studies and/or from differences in the effects of management on vertical accretion and soil among impoundments. One major difference is the avoidance of managed drawdowns in Unit 15, and the inability to drawdown Unit 6 (for agricultural reasons), throughout the existence of both of those units. Drawdowns were rarely used in Unit 2 prior to 2005 when it was managed to maximize alligator nest density, but it was frequently managed with drawdowns after ~ 2007 when it was managed to maximize use by migrating shorebirds, wading birds, and waterfowl. It is widely recognized that soil drainage accelerates soil organic decomposition. For example, Nyman and DeLaune (1991) showed that lowering the water level one centimeter below the soil surface accelerates CO₂ emissions 4% in soils dominated by Panicum hemitomon, which dominates many fresh marshes in southeast Louisiana, and in soils dominated by Spartina patens, which dominates intermediate and brackish marshes in Louisiana, even though CO₂ emissions differ drastically between those soils. Each additional centimeter of drainage similarly increases CO₂ emissions Nyman and DeLaune 1991). In addition to accelerated decomposition of soil organic matter, drainage is widely recognized to compact organic soils and thus increase soil bulk density. This is the most likely reason that Unit 2 had greater soil bulk density (0.46, SE = 0.02) than the un-impounded marsh (0.22 15 g cm³, SE = 0.02). It is less likely that the adoption of drawdowns not only prevented vertical accretion after 2005, but also negated much of the elevation gained between 1963 and 2005.

Management Implications

Effects of impoundment on accretion can vary widely, which precludes broad statements about effects of impoundment on the ability of coastal marshes to accrete. However, our findings agree with those of (Drexler et al. 2013) who also concluded that moist-soil management on organic soils in the coastal zone compromises accretion processes and can lead to oxidation of soils and loss of elevation. Many marshes can keep up with sea level rise through vertical accretion processes, but our results indicate that their ability to do so can be compromised by moist soil management. Furthermore, if organic soils are oxidized and sites lose elevation, levees can be compromised during high water levels (Bates and Lund 2013) and dramatically increase the costs of management.

Despite these challenges, however, water level management will likely remain an important management strategy in these regions because they provide tremendous habitat resources in a small area. Furthermore, with the continued deterioration of surrounding marshes (Coastal Protection and Restoration Authority 2023), the habitat values provided by these impoundments remain an important component of waterbird conservation strategies. A process-based understanding of specific management strategies, such as salinity pulses and timing and duration of drawdown on both productivity and decomposition would inform management decisions. Although the specific impacts of many processes are unknown, it is established that some strategies can reduce overall impacts of management. For example, maintaining high groundwater levels so that underlying organic soils remain saturated can reduce peat decomposition (Kechavarzi et al. 2007; Boonman et al. 2022). Managers will need to understand how seed germination of desirable plants are affected by soil moisture dynamics from high water tables, but this should reduce elevation loss because of the anaerobic conditions in the saturated zone. Another potential strategy is to promote accretion during some years and manage for mudflat annuals in others. For example, implementing Open Marsh Water Management (OMWM) to offset artificial drainage caused by mosquito ditches by plugging ditches and raising the water table has increased vertical accretion by 30% and carbon accumulation by 23% over that in marshes with minimal disturbance from human activities such as dredging, filling, ditching, or dominance by invasive species (Drake et al. 2015; Table 4). That study was based on data from four National Wildlife Refuges along the New England coast of the USA (Drake et al. 2015). Two other examples lie in the Sacramento-San Joaquin Delta of California, USA. There, wetlands were allowed to develop on two permanently flooded 3-ha impoundments excavated in a subsided agricultural field in 1997. Accretion in those impoundments averaged ranged from 2.4 cm yr^-1 to 3.6 cm yr^-1 between 1997 and 2017 (Miller et al. 2008; Deverel et al. 2020). That elevation gain was controlled by organic matter accumulation and was accompanied by ~ 1 kg of carbon storage (Miller et al. 2008). In 2014, agriculture ceased on the surrounding agricultural field. There, accretion averaged 3 cm yr^-1 between 2014 and 2018 (Deveral et al. 2020). All three examples resulted from managers who raised water levels in wetlands or former wetlands where water levels were artificially low. A managed marsh in coastal Massachusetts, USA provides a fourth example of wetland managers accelerating accretion. There, Eagle et al. (2022) found that restoring tidal exchange to previously impounded coastal marshes allowed for an accretion rate of over 13 mm/year. These four examples suggest that management for accretion can lead to rapid results and suggests that to increase accretion and manage units for waterfowl, coastal wetland managers could alternate between several years of lowering water levels to prioritize waterbird use, and several years of shallow flooding to prioritize accretion via vegetative growth. This suggestion is untested, however, and warrants future study.

Acknowledgements

We thank the Louisiana Department of Wildlife and Fisheries and Rockefeller Refuge for financial and logistical support. A. Booth, A. Beaudette, M. Gill, P. Trosclair, J. Marty, G. Giffin, and C. Wilson provided invaluable field and technical assistance. Any use of trade, firm, or product names is for descriptive purposes only and does not imply endorsement by the U.S. Government. J.A. Nyman’s participation and contribution were based upon work that was supported by the National Institute of Food and Agriculture, US Department of Agriculture,

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Managing Marshes with Drawdowns Can Reduce Elevation of Coastal Marshes

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