9/25/2014 TEXAS COASTAL E XCHANGE ECOSYSTEM SERVICES OF THE MID-­‐TEXAS COAST Prepared for the SSPEED Center: A review of available literature on the metrics and values of services provided by coastal ecosystems | Courtney Hale, Avantika Gori, Jim Blackburn Contents List of Tables and Figures .......................................................................................................................... 2 Methods for Valuing Ecosystem Services ................................................................................................. 4 Oyster Reefs .................................................................................................................................................. 6 Oyster Production ..................................................................................................................................... 6 Augmented Fishery Habitat ...................................................................................................................... 8 Water Quality .......................................................................................................................................... 11 Breakwaters/Erosion Protection ............................................................................................................. 14 Non-­‐Use Value ........................................................................................................................................ 16 References .......................................................................................................................................... 17 Coastal Wetlands ........................................................................................................................................ 19 Wetland Services ................................................................................................................................. 19 Fishery Support ....................................................................................................................................... 20 Bird Support ............................................................................................................................................ 22 Carbon Sequestration ............................................................................................................................. 24 Storm Protection ..................................................................................................................................... 26 References .......................................................................................................................................... 28 Coastal Bottomland Hardwood Forests ...................................................................................................... 31 Columbia Bottomlands ....................................................................................................................... 31 Ecosystem Services of a Bottomland Hardwood Forest ..................................................................... 32 Inundation Capacity ................................................................................................................................ 33 Nutrient and Pollutant Reduction from Runoff ...................................................................................... 36 Carbon Sequestration and Sink Capacity ................................................................................................ 40 Tree species composition.................................................................................................................... 40 Carbon accumulation and stock .......................................................................................................... 41 Migratory Bird Habitat and the Birding Industry .................................................................................... 43 References .......................................................................................................................................... 46 Texas Coastal Prairies.................................................................................................................................. 50 Carbon Sequestration ............................................................................................................................. 50 Nutrient Cycling ...................................................................................................................................... 52 Bird Habitat Support ............................................................................................................................... 54 References .......................................................................................................................................... 55 1 List of Tables and Figures Table 1: Services provided by the ecosystems of the Texas coast ............................................................... 3 Table 2: Historic and Present Areas of Texas Oyster Reefs .......................................................................... 7 Table 3: Oyster Reef Fishery Values Summary Table .................................................................................. 11 Table 4: Oyster Reef Water Quality services Summary Table .................................................................... 14 Table 5: Oyster Reef shore protection value summary table ..................................................................... 16 Table 6: Wetlands Fishery Support Summary Table ................................................................................... 22 Table 7: Wetlands Carbon Services Summary Support .............................................................................. 26 Table 8: Bottomland Forest Nutrient Removal/Retention Summary Table ............................................... 40 Table 9: Bottomland Forest Carbon Services Summary Table .................................................................... 43 Table 10: Bottomland Forest Bird Habitat Values ...................................................................................... 45 Table 11: Prairie Carbon Services ............................................................................................................... 52 Figure 1: Some functions provided by oyster reefs ...................................................................................... 6 Figure 2: Shucking an oyster harvest in Beaumont, TX................................................................................. 8 Figure 3: An oyster reef provides foraging sites for fish ............................................................................... 9 Figure 4: Water quality at a site in Chesapeake Bay before and after an oyster reef restoration ............. 12 Figure 5: Oyster reef acting as a breakwater to protect coastal marsh in Alabama .................................. 15 Figure 6: Services provided by a coastal wetland ....................................................................................... 19 Figure 7: Terracing in Galveston Island State Park as seen from Google Earth .......................................... 21 Figure 8: White ibis flock over a coastal wetland ....................................................................................... 23 Figure 9: Marsh transition towards upland habitat .................................................................................... 24 Figure 10: Columbia Bottomlands Conservation Area ................................................................................ 31 Figure 11: Typical functions of a bottomland wetland forest (Sun 2002) .................................................. 32 Figure 12: About half of the area of the Bottomlands falls below 25 foot elevation ................................. 34 Figure 13: The change in peak flows when a forested watershed is converted to urban land .................. 35 Figure 14: Hypoxic zones shown in red along the coasts of Louisiana and parts of Texas ........................ 36 Figure 15: Inundated area of Brazos Bend State Park ............................................................................... 37 Figure 16: Neotropical migration routes intersecting over the Columbia Bottomlands ............................ 44 Figure 17: Beautiful birds such as this Baltimore Oriole can be seen each year as they migrate through the Columbia Bottomlands ......................................................................................................................... 46 Figure 18: Root systems of common prairie plants .................................................................................... 50 Figure 19: A sparrow and a flycatcher are two common birds to see on a prairie grassland .................... 54 2 The Texas Gulf Coast is home to numerous ecosystems that provide humans with an abundance of different services. Ranging from the bay itself, oyster reefs and estuaries, prairies, bottomland forests, to wetlands, the Texas Gulf Coast supports a diverse group of habitats, and thus proves to be invaluable to the state of Texas. However, in recent decades much of these precious lands have been destroyed or degraded due to land conversion, anthropogenic pollution, over-­‐use of ecosystem resources, and other factors. This paper aims to review the functions of well-­‐known ecosystems along the Texas coast to better understand the magnitude at which services are provided to us and their potential economic values. /Ŷϭϵϵϳ͕ZŽďĞƌƚŽƐƚĂŶnjĂĐŽŶƚƌŽǀĞƌƐŝĂůůLJĚĞĐůĂƌĞĚƚŚĂƚƚŚĞǀĂůƵĞŽĨƚŚĞǁŽƌůĚ͛ƐĞĐŽƐLJƐƚĞŵƐĐŽƵůĚďĞ
estimated at $33 trillion per year in 1995 $US (Costanza et al, 1997). CostannjĂ͛ƐǀĂůƵĂƚŝŽŶƉƵƐŚĞĚƚŚĞ
scientific community into further investigation of the value of ecosystem services and has since inspired numerous studies and economic analyses on the function services provided by natural ecosystems. Earlier this year, Costanza released an updated version of his original paper, documenting changes to ƚŚĞǁŽƌůĚ͛ƐĞĐŽƐLJƐƚĞŵƐ͕ĂŶĚĞdžƉůĂŝŶŝŶŐŚŽǁƚŚĞƐĞĐŚĂŶŐĞƐŚĂǀĞŝŶĐƌĞĂƐĞĚƚŚĞǀĂůƵĞŽĨĚŝĨĨĞƌĞŶƚ
ecosystems; it is evident in his analysis that has ecosystems become further degraded and developed, the sacristy of their services increases the service value tremendously. The Millennium Ecosystem Assessment of 2005 classifies ecosystem services into four categories: provisioning, regulating, supporting, and cultural. Ecosystems on the Texas coast provide services that fall into all of these broad categories. Provisioning services include the product harvest from an ecosystem such as oysters, fish, or timber. Regulating services include water quality maintenance, shore protection, carbon services and shore stabilization. Nutrient cycling, aquifer recharge services and providing nursery habitat for other species are considered supporting services, and cultural services are provided by the recreation and tourism associated with ecosystems and the historical and aesthetic qualities these ecosystem types yield. The ecosystem services provided by four ecosystems of the Texas coast are shown in Table 1. T ABLE 1: SERVICES PROVIDED BY THE ECOSYSTEMS OF THE TEXAS COAST Provisioning ƒ Oyster harvest ƒ Shell material Oyster Reefs Coastal Marshes Prairies Bottomland Forests ƒ Crab and shrimp harvest ƒ Grazing land for ranching ƒ Hunting land ƒ Timber harvest ƒ Hunting land Regulating ƒ Water quality/filtration ƒ Shore erosion control ƒ
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Supporting Cultural/Non-­‐Use ƒ Fishery habitat ƒ Estuarine protection ƒ Aquifer recharge ƒ Fishery estuary Habitat ƒ Recreation ƒ Historical significance Carbon Sequestration Flood Protection ƒ Aquifer recharge ƒ Bird Habitat ƒ Aesthetic Beauty Carbon sequestration Air Quality Water quality Denitrification ƒ Bird Habitat ƒ Wildlife corridors ƒ Recharge aquifers ƒ Aesthetic beauty ƒ Recreation ƒ Bird-­‐watching areas Carbon Sequestration Water quality Wave attenuation Surge protection ƒ Recreation ƒ Aesthetic 3 Methods for Valuing Ecosystem Services Different methods have been developed with varying levels of acceptance to measure a services value to humans monetarily. Provisioning services are relatively easy to value because a market exists where values can be directly monitored when products are bought and sold; changes in particular ecosystem can also be attributed to changes in values of the provisions if the market is regulated thoroughly (Northern Economics 2009). Regulating services are considered indirect services. The cost of a substitute of an ecosystem function can be a valid method to determine the value of an indirect service (Henderson 2003). For example: the cost to protect a shoreline using engineered techniques as opposed to an oyster reef can be directly transferred as the value of the shore protection services provided by an oyster reef. The cost to filter water to obtain similar nutrient retention as provided by a wetland or bottomland forest could also be directly transferred as the value of that ecosystem service. As markets for nutrient trading begin to emerge due to water quality issues in surface water bodies, the market value of nutrient removal by ecosystems may be more easily evaluated (Northern Economics 2009). The supporting services that certain ecosystems provide to other invertebrate and fish species, in addition to those species that are commercially fishable, can be estimated using additional population and density estimates that are converted into commercial market values for each species group: game fish, bait fish, crab, etc. However, this approach would not be able to capture a market value for species that are not commercially valuable, but may be ecologically valuable. It is important to note the differences in the population density of harvested species before and after an oyster reef restoration or wetland marsh restoration activity to understand the augmented value that that reef provides to the fishery species. The credited value attributed to oyster reef can only account for the augmented fishery biomass value, not the entire production value as some biomass may have existed prior-­‐to and regardless-­‐of restoration of the reef. Cultural services are the non-­‐tangible benefits that people derive from the ecosystem such as recreation, aesthetic enhancements, spiritual enrichment, and historical value. The historical landscape of the Texas coast has had a significant role in the development and history of coastal areas and communities. Valuing the coastal services of an ecosystem can really only be accomplished by gathering information from the people who value the ecosystem in this way. Travel time and money spent to engage in activities reliant on specific ecosystems or willingness to pay for restoration or preservation for aesthetic or historical benefits are decent ways to gauge the cultural value of an ecosystem. Because the cultural services provided by oyster reefs are not tangible and cannot be directly measured, they are the most difficult and ambiguous services to quantify. There are acceptable and practiced ways in which economists place monetary values on specific ecosystem services. However, the task of applying monetary values to whole ecosystems can be very arduous. The replacement cost or substitute cost method only accounts for a single service, and often an ecosystem provides many services through various functionalities that interact with each other, making the holistic valuation of an ecosystem very difficult and error-­‐prone. The benefits estimated using these described methods of replacement cost, augmented commercial value, cost-­‐transfer, and market value cannot simply be added together to form a complete total value of ecosystem services. 4 Some benefits may overlap or influence others, thereby increasing the risk of double-­‐counting benefits by adding values. For example, the filtering capacity of an oyster reef enhances water quality which is a service that has a specific value, but enhanced water quality may be an additional factor that attracts certain species to the reef rather than the attracting force being just the presence of the reef. Therefore the improved water quality value may overlap with the augmented fishery value. Interdependent functionalities, such as these, are what make economic valuation of holistic ecosystems so difficult. 5 Oyster Reefs Oysters have long been recognized as a commercial and recreational harvestable economic good, and the Texas gulf coast is the second-­‐largest producer of oysters in the nation. Shellfish and oyster habitat is one of the most threatened ecosystems on Earth, with an estimated 85% loss having occurred historically throughout the world. Dwindling oyster populations and continued threats on this declining ecosystem is a causing concern for the potential losses of ecosystem services that are offered by oyster reef habitat. Oysters and their reefs offer many ecological services, in addition to acting as a consumer good, that we as humans benefit from including providing critical habitat for juvenile fishery species, the ability to filter large volumes of water to improve water quality and nutrient cycling within bays and estuaries, acting as breakwaters to reduce shore erosion and providing stable sediment areas that increase aquatic vegetated zones which further improve water quality and commercial fishery estuarine habitat (Figure 1). Quantifying the functional value and economic benefits of oyster reefs is important to understand the extent of services provided, and to justify restoration efforts to improve or reestablish the lost functionality (Beseres Pollack 2013) While traditional oyster reef or estuarine studies have simply aimed to prove the existence of ecosystem service functions, numerous studies have surfaced over recent years that attempt to quantify oyster reef ecosystem services in terms of ecosystem service functionality and economic value. adapted from: Using the described valuation methods, http://sites.psu.edu/ank5283/2013/10/02/chesapeake-­‐
disconnect/ many studies both quantify the ecosystem F IGURE 1: SOME FUNCTIONS PROVIDED BY OYSTER REEFS functionality of oyster reefs for a specific service and then convert those values into dollars. There are four major ecosystem services provided by oyster reefs that tend to hold the most economic value and are able to be quantified rather directly. This section will summarize the quantities and values from several different studies for the services of oyster production, augmented fishery value and habitat, shoreline stabilization and erosion control, and filtration of nutrients for water quality enhancement. Oyster Production The most obvious service provided by oyster reefs is the harvest of the oysters themselves. Oysters are harvested and sold commercially as a consumer good; they are one of the only organisms that are widely consumed alive and raw (Figure 2). Texas Parks and Wildlife reports that over 6.1 million pounds of oyster meat, valued at over $11.1 million, were pulled from and commercially sold out of Texas bays 6 during the 2000 season (Robinson 2014). Even though this is an astounding volume, and Texas has seen even larger volumes of oyster harvest since 2000, overharvesting has dramatically reduced oyster populations and production over the last several decades. An analysis by Zu Ermgassen et al (2012) concluded that there has been a 64% decline in oyster reef extent and 88% loss in oyster biomass between early 1900s and 2000s for the entire United States. The historic area of oyster reef habitat in Texas bays has been estimated based on historic data and current maps to have been depleted by an average of 40%, with Galveston Bay͛Ɛ ĂŶĚ^ĂŶŶƚŽŶŝŽĂLJ͛ƐŽLJƐƚĞƌŚĂďŝƚĂƚďĞŝŶŐdepleted 83% in the last century (see Table 2). T ABLE 2: HISTORIC AND PRESENT AREAS OF TEXAS OYSTER REEFS Galveston Bay Matagorda Bay San Antonio Bay Aransas Bay Corpus Christi Bay Total Historic Area (acres) 32000 41215 6400 9600 8320 Present Area (acres) 26675 5508 5332 1191 716 Percent Loss 83.4 13.4 83.3 12.4 8.6 97535 39422 40.4 Source: adapted from Zu Ermgassen et al (2012) Calculating the per acre harvest value of Texas oysters using the 2000 value from TPWD and the reported harvestable acres of oyster reefs of 25,081 acres, the annual value of oyster harvest from harvestable oyster reefs in Texas bays is a state-­‐wide average estimate of 243.2 pounds of meat/acre or $442.56/acre (Robinson 2014). 1 Grabowski et al (2012) observed the value of oysters harvested from pristine reefs is significantly greater than the value of those harvested from degraded reefs, both in terms of density and quality. The same study also noted that destructive harvesting techniques, such as dredging, greatly reduced the overall value of an oyster reef and the value of the harvested specimens over time, preventing a restored reef from recovering initial restoration costs. A study conducted for the Corp of Engineers looked at two oyster reef restoration projects on the east coast and estimated the cost of construction at $10,000/acre and brood stocking the acre of reef as an additional $10,000. When evaluated for the value of harvesting the oysters, it was estimated that the years it would take to recover the initial cost of the reef restoration was between 7-­‐14 years without annual maintenance and a harvest density of 100-­‐200 bushels per acre. In the seeded reef, it was estimated it would only take 2-­‐5 years to recover initial costs 1
This calculation does not account for varying locations, quality of oysters, densities of oysters, or harvest accessibility of the numerous individual reefs that comprise the total harvestable acres, and only represents one ƐŽƵƌĐĞ͛ƐƌĞƉŽƌƚ͘ 7 at that same harvest density (Henderson 2003). However, this analysis was only considering the harvest value of the reefs for recovery value and times; oyster reefs contribute many other ecological services to the surrounding waters and habitats of the area they are restored in, contributing many other significant economic values. The time it take a resorted reef to recover initial costs in total value may be drastically lower if all other economic values of services are considered in addition to harvest value. Valuing the production service of oyster reefs for harvesting oyster meat is simple process. The net profit acquired by the oystermen is a direct transfer to the value of the oysters. There is also additional profit made through the resale of oysters first to distributors, to restaurants, and then to consumers that can be added to the initial dock value. It is a simple methodology, yet can become very tedious to pinpoint a single service value due to the tremendous variability of harvest from reef to reef, bay to bay, and season to season. Different reefs and bays may have an average value per pound of meat, per acre of oyster reef, or per bag of whole oysters that is reported different depending on the season, location, or current market price; comparing and aggregating these varying values can make it difficult to come up with a total value for the entire oyster reef system unless every transaction were to be monitored effectively and accurately. Because of http://ww2.hdnux.com/photos/06/37/65/1703237/3/628x471.jpg recreational and subsistence harvesting, and the chance of F IGURE 2: SHUCKING AN OYSTER HARVEST IN BEAUMONT , TX unregulated or under-­‐the-­‐table transactions, the true value of oyster harvest is likely much higher than what is traditionally reported as commercial harvest value. Augmented Fishery Habitat Oyster reefs have been seen to provide protective habitat to shrimp when a predatory species is detected; it was observed that the shrimp in a study by Coen et al (2000) do not use the reef as protection when a non-­‐predatory species is present. Finfish use the reef in three ways: as a primary habitat, intermittently as protection or shelter, or for transient foraging (Figure 3). Additionally, resident species utilize microhabitats found in the spaces between oyster shells to lay eggs or as nesting sites (Coen 200). A study by Zimmerman et al (1989) assessed an oyster reef habitat in West Bay, Galveston, Texas, for the presence and abundance of several economically valuable species such as grass shrimp, blue crab, stone crab, and several game fish species. The densities of the fishery species was compared on an oyster reef and in a salt marsh and bare mud flats using drop trap sampling. This study showed that winter and summer fish abundances were higher at the oyster reef than at the other two sample habitats. Stone crab was also shown to be more abundant at the reef in both seasons. Overall, the reef 8 and marsh habitat were significantly more used than the mud flat habitat. Both the reef and salt marsh had unique species assemblages and are therefore not considered to be interchangeable habitat (Zimmerman 1989). Oyster reef habitat enhances recruitment of juvenile stages of fishery species and the growth and survival of these species by providing food, shelter, and protection from predators. At the landscape-­‐
scale, oyster reef restoration is a feasible way to increase fish habitat and consequently increase the production and resilience of fishery systems in the Gulf (Grabowski 2012). An aggregation of data from 6 studies conducted in southeast USA estimated that 10 sq. meters of restored oyster reef habitat creates an additional 2.6 kg of fish and large crustacean production per year (Peterson 2003). This is equivalent to 2319.6 pounds of fish production per acre of reef per year (lbs./ac/yr.). Two studies conducted in North Carolina in 1997 by North Carolina Sea Grant on the value of augmented fishery production by oyster restoration observed that the long-­‐term value of fish and crab production for commercial harvest was greater than the value of the commercial harvest of oysters from the reefs. The studies also concluded that the value of fish caught on a restored reef and the value of those caught on a natural reef was the same (North Carolina Sea Grant 1997). In Louisiana, recreational anglers fishing over oyster beds, because of improved water quality and habitat of the area, resulted in an estimated economic benefit to the state͛s coast of $2 Million (Henderson 2003). The irregular surface that is created by an oyster reef holds 50 times the surface area than the same-­‐sized flat bottom habitat, providing protection and nursery sites for many different species of invertebrates and small juvenile or bait fishes; this is what makes oyster beds and reefs such an attractive fishing location. (Henderson 2003). In Louisiana and other estuaries along the Gulf of Mexico, the narrow tidal range prevents many oyster reefs from growing a three-­‐
dimensional structure and instead results in large, flat, subtidal oyster beds. A http://www.cbf.org/image/area-­‐-­‐-­‐about-­‐cbf/offices-­‐operations/virginia/fish-­‐checks-­‐out-­‐oyster-­‐
study by Plunket in 2005 reef_image008_CBFStaff.jpg compared the fish and crab F IGURE 3: A N OYSTER REEF PROVIDES FORAGING SITES FOR FISH abundance and diversity in these subtidal oyster beds to mud bottoms of the same bay. It was shown that the flat oyster bed structure seen in bays around the Gulf are more diverse and densely populated by crustaceans, mussels, and bottom feeding fish than mud bottom. The densities of catches over the oyster beds was around 200-­‐250 catches/m2 where the mud bottom density of catches was around 150 per m2 with the majority of the difference coming from the catch densities of crustaceans, mud bottoms having approximately half the density of the oyster 9 beds. Total numbers of fish caught and number of species caught by gill nets were both greater over an oyster bed habitat than the mud bottom (Plunket 2004). Shell deposits in Louisiana, assumed to also occur in other gulf coastal areas, and other reefs restored for oyster production also provide habitat for fish and decapod crustaceans. The benthic animals that reside in these cultivated oyster beds are valuable to foraging transient fish. This study conducted by Plunket concluded that providing habitat such as a cultched oyster bed is likely to enhance the value and commercial productivity of the entire surrounding estuarine ecosystem (Plunket 2004). A study by Stunz et al in 2010 conducted in West Galveston Bay, Texas, consisted of shallow open water, fragmented salt marsh, and oyster reefs. The highest densities of fishes occurred on the oyster reef sites and the highest decapod densities occurred on the oyster reefs and marsh edges. The oyster reef sites had a mean density of 17.22 /m^2, marsh edge was 4.02/m^2, and bottom was 7.6/m^2 and 6.54/m^2 for near and far respectively. The crustacean mean densities are 62.27/m^2 for oyster reef, 55.03/m^2 in marsh edge, and 5.05/m^2 and 4.08/m^2 for bottom near and far. The oyster reef sites also supported the greatest number for species richness in all seasons (Stunz 2010). A study near Mobile Bay, Alabama looked at two oyster reefs designed as breakwaters to analyze the augmented fishery recruitment. Using gillnets and seines along the edges of the reef, fish and crustacean biomass and species diversity was recorded and compared to control sites that had mud bottoms. The results showed a strong enhancement of demersal fish species near the reefs, with spotted sea trout 38% more abundant near reefs. The species richness was significantly higher near the reefs from the 10 cm net catches. Decapod crustaceans were more abundant and had greater densities at the reef sites than on the mud bottoms. This study showed that there was a significant relationship ďĞƚǁĞĞŶƚŚĞƉƌĞƐĞŶĐĞŽĨĂŶŽLJƐƚĞƌƌĞĞĨ͕ĂŶĚĨŝƐŚĂŶĚĐƌƵƐƚĂĐĞĂŶĂďƵŶĚĂŶĐĞ͕ďƵƚĚŝĚŶ͛ƚŐŽĂƐĨĂƌĂƐ
quantifying the augmentation value (Scyphers 2011). An analysis by Timm Kroeger for The Nature Conservancy in 2012 looked at two oyster reef restoration projects in the Gulf of Mexico and computed quantitative functional values for ecosystem service provided by those oyster reefs. The augmented fishery value for this study was based on findings from literature. The denitrification values were also pulled from literature sources. The wave attenuation estimates were generated using a standard hydraulic model of the reefs. Using data from both Peterson, 2003 and Scyphers, 2011, as well as additional studies from the northern Gulf of Mexico, Kroeger aggregates fishery augmentation data into a complete total value of fish for the total length of the two reefs in the study, 3.6 miles. Kroeger adjusted values from the literature to apply to only fishes that were sizable enough to be of interest to fisherman and therefore, market available. He concluded that the reefs of his study will result in an increased production of the species included in the studies of 296 g/m2 of reef/year, or 2640.8 pounds of fish/acre of reef/year. The harvestable production increase was estimated to be 58% of the total increase of production, based on different fish sizes and age classes accepted commercially, which is calculated from the total augmented production to be 1531.6 pounds of harvest/acre of reef/year (Kroeger 2012). 10 T ABLE 3: OYSTER REEF F ISHERY VALUES SUMMARY TABLE Fishery Production and Augmentation Source Location Values Peterson (2003) Southeast USA 2319.6 additional lbs. fish and crab production/acre/year Henderson (2003) Louisiana $2 million benefit from recreational anglers fishing over oyster reefs Plunket (2005) Louisiana 200-­‐250 organisms/sq. m of reef Stunz (2010) West Bay, TX 17 fish/sq. m of oyster reef 62 crustaceans/sq. m of reef Scyphers (2011) Mobile Bay, AL 38% increase in sea trout population Kroeger (2012) Mobile Bay, AL 2640.8 lbs. augmented population/acre/yr. 1531.6 pounds of harvest/acre/year AVERAGES FROM ALL STUDIES: 2480 lbs. production/acre/year 176 organisms/sq. m of reef Water Quality Large congregations of oysters and other bivalves associated with oyster reefs have a great impact on the water quality of the bays and estuaries to which they reside. As filter feeders, oysters and other bivalves have the ability to remove sediment and nutrients from the water column at surprising rates, contributing to the overall nutrient cycling, carbon sequestration, nitrogen control, and reduction of algae blooms of our bays and coastal gulf waters. In addition to dissolved nutrients, phytoplankton and suspended solids are removed through the process which increases water clarity and light penetration into shallow water, therefore impacting sea grass beds and submerged aquatic vegetation (SAV) areas that further improve water quality and act as valuable nursery grounds for many fishery species (Zu Ermgassen 2013).2 Oysters filter the water by removing particles and nutrients from the water column, digesting the edible components, assimilating into shells or tissue, and depositing other components onto the sediment or onto the reef structure (Figure 4). Several in situ water conditions and factors as 2
The value of the additional SAV fishery and estuarine habitat created could also be credited to oyster reefs, although there are not any current studies that quantify this value based on the indirect impact from improved water quality, but rather only the density of fish species related to the oyster reef itself. 11 well as size of the oyster directly affect the volume of water filtered and the efficiency of filtration (Zu Ermgassen 2013). It has been well demonstrated through both laboratory study and field observation that suspension-­‐feeding shellfish such as oysters have impact on a basin-­‐wide water quality and phytoplankton activity (Coen 2000). Several Texas estuaries were looked at in a study by Zu Ermgassen in 2013 and analyzed to determine the historical and present filtration capacity of oysters. The results show that currently, oysters in Galveston Bay filter 210 m3 of water per ha per hour at their present density. Results for other bays include the following: 220 m3/ha in Matagorda Bay, 210 m3/ha in San Antonio Bay, 790 m3/ha in Aransas Bay and 70 m3/ha in Corpus Christi Bay. These values are dramatically less than those seen on the eastern coast estuaries and represent a 97-­‐100% loss of filtration capacity compared to historical levels. The value of filtration ability of oysters is increasing because of the high rates of eutrophication in many shallow estuary systems, yet the capacity of our oyster populations to fulfill this role is diminishing. The reduction of filtration http://www.pbs.org/newshour/rundown/restoring-­‐the-­‐gold-­‐of-­‐the-­‐chesapeake-­‐
capacity due to the bay/ dramatic decline in oyster densities and extents is likely to be the cause of substantial changes in the ecosystem functions of these estuaries. It is unlikely that restoring oyster beds closer to the historically functional densities would F IGURE 4: WATER QUALITY AT A SITE IN C HESAPEAKE B AY BEFORE AND AFTER AN completely solve our OYSTER REEF RESTORATION water quality issues, but doing so could cause a significant increase in the overall health and functionality of these estuarine ecosystems, on a local and landscape scale (Zu Ermgassen 2013). Oysters filter nitrogen out of coastal waters, helping reduce the levels that cause over enrichment and eutrophication of these areas, a large global issue. Traditionally, anthropogenic nutrient load control has been the go to strategy for dealing with over enrichment issues, but bivalve filtration and sequestration provided by increasing volume of oyster reef habitat may be a way to help solve this global crisis and avoid further regulatory controls (Beseres Pollack 2013). Bivalves remove nitrogen from the water column in the form of planktonic nitrogen and can assimilate an estimated 50% of particulate organic nitrogen (Beseres Pollack 2013). This nitrogen is removed with oyster harvest or sequestered into shell and tissue and available to tertiary consumers of the oyster. The other 50% is biodeposited and either 12 buried or undergoes denitrification, both processes that result in nitrogen removal from the water column (Beseres Pollack 2013). A study by Beseres Pollack in the Mission-­‐Aransas estuary of Texas calculated through their methods that the oysters in those bays comprised 18.11 square km of reef and removed 35,315 kg of nitrogen in the year of study. This is equivalent to a computed rate of 17.4 lbs. N/acre/year. Their cost-­‐transfer analyses concluded that removing that amount of nitrogen using a water treatment facility constituted a value of nearly $294,000 annually (Pollack 2005). When oysters are removed from an ecosystem, nitrogen is removed with them, but the capacity for further filtering of the water is also diminished. This creates a trade-­‐off between once time harvest value and the value of continuing ecosystem services provided by leaving the oysters in place (Beseres Pollack 2013). Grabowski et al in 2012 used denitrification rates from Piehler and Smyth (2011) which concluded from a study conducted in North Carolina that the net augmented denitrification rate of an oyster reef is 143 Kg N/hectare/yr., or a computed 127 lbs. N/acre/yr. (Piehler and Smyth 2011). Grabowski used a different type of cost replacement approach and utilized the average trading price for a kilogram of nitrogen removal in the North Carolina Nutrient Offset Credit Program of $28.23 (likely to rise) to compute a final value of nitrogen removal from 1 ha of reef to be estimated at $4050/yr. in 2011 dollars, or $1640/acre of reef. Using a model developed by Zu Ermgassen, whose studies and report is discussed above, and data from two studies on denitrification rates of oysters on the east coast, converted to proper temperatures and salinity conditions found in the Gulf of Mexico, Kroeger 2012 estimated the annual nitrogen removal rate of the oyster reefs in his study in the northern Gulf of Mexico to be 0.14-­‐2.81 kg N/ha/day, or 45-­‐
915 pounds N/acre/yr. (Kroeger 2012). Because the model used in this study was developed using the only two studies available that document results from field experiments, and both of the studies were conducted on the east coast, it is difficult to get an accurate assessment of denitrification rates in the Gulf of Mexico without conducting similar field studies. However, from the combination of information from all of the studies looked at for this report, it is possible to gain some information on the order of magnitude and potential range of rates of denitrification that occur along the Texas Gulf Coast. 13 T ABLE 4: OYSTER REEF WATER QUALITY SERVICES SUMMARY TABLE Water Quality-­‐ Nitrogen Removal and Total Filtration Source Location Values Beseres Pollack (2013) Mission Aransas, TX 17.lbs.bs N/acre of reef/year Kroeger (2012) Mobile Bay, AL 45-­‐915 lbs. N/acre/year (winter-­‐summer) Piehler (2011) North Carolina 127 lbs. N/acre/year Grabowski (2012) North Carolina $1640 of Nitrogen removal/acre/year AVERAGE FROM ALL STUDIES: 276 lbs. N/acre/year Zu Ermgassen (2013) Galveston Bay, TX 22450.39 gallons water/acre/hour Matagorda Bay, TX 23519.46 gallons water/acre/hour San Antonio Bay, TX 22450.39 gallons water/acre/hour Aransas Bay, TX 84456.24 gallons water/acre/hour Corpus Christi Bay, TX 7483.46 gallons water/acre/hour AVERAGE FROM ALL STUDIES: 32072 gallons/acre/hour Breakwaters/Erosion Protection Near-­‐shore estuaries and vegetated habitat represent 23.7% of global ecosystem services (Scyphers 2011). The ecosystem services provided by these vegetated estuaries and shoreline marsh habitats are coming under increasing threat from predicted sea level rise, increased utilization, and sensitivity to severe storms and damaging wave energies. Engineered structures designed to mitigate these issues such as bulkheads and seawalls are often used with no regard to their ecological effects on the greater intertidal and benthic ecosystems connected to the shoreline (Scyphers 2011). In tidal and subtidal environments, an oyster reef stabilizes sediments and deflects wave energy that protects shorelines and prevents erosion (Henderson 2003). The stable sediment also allows for the establishment of submerged aquatic vegetation. Loss of shoreline and shore erosion has been attributed to decreasing coastal property values. Because oyster reefs also have the ability to promote sedimentation, shorelines can actually be altered or created, providing more valuable coastal marsh or inland habitat (Henderson 2003). Because oyster reefs provide these shore protection services while simultaneously providing other ecosystem services, improving instead of damaging existing intertidal ecosystems, oyster reefs are a great alternative to engineered shoreline protection systems (Scyphers 2011). 14 Oyster reefs can function just like man-­‐made jetties or bulkheads because they interact with wave energy the same way, by buffering the wave and increasing sedimentation (Grabowski 2012). Engineered shoreline stabilization devices costs could be used instead to restore oyster reef along the shoreline, at equivalent functionality to the land-­‐owner. Because oyster reefs can grow vertically with sea-­‐level rise, they can be attributed a higher value than engineered devices because they are naturally resistant to sea level rise and are able to regenerate themselves following any damaging events (Grabowski 2012). The location of restoration activities is very important in determining the cost value of erosion control services provided by the reef. In areas where engineered systems would traditionally be built, the value is equivalent to those systems, discounting the fact that the reef also provided additional services such as those discussed in this paper. If the reef is built in an area where erosion is not traditionally a concern, then the value of the reef, strictly in terms of erosion protection services, may be evaluated as very low or zero, simply because the services are not needed and not marketable (Grabowski 2012). A study conducted in Mobile Bay of Alabama looked at mitigating shoreline loss with oyster shell breakwaters at two different locations along the gulf facing shore and inner-­‐bay shore (Scyphers 2011). The shoreline and bathymetry observations showed that the oyster shell reef decreased the depth of water between the breakwater and the shore and increased sedimentation on shore, expanding the footprint of the shoreline 300% compared to the control site (Scyphers 2011). The study also found that breakwater reefs constructed using loose oyster shell encouraged oyster recruitment and held a greater diversity of fishes and invertebrates than areas with no constructed reef (Scyphers 2011). These benefits are not seen in traditional shore-­‐armoring techniques. In areas where defenses are already present or immediately necessary, providing oyster shell breakwaters in addition to shore retention engineering would mitigate the loss of the fish and crustacean habitat caused by construction of the engineered walls (Scyphers 2011). Another study conducted in Louisiana compared the potential shoreline protection of oyster reefs in a high wave energy environment and a low energy environment (Piazza 2005). The results of this study showed that shoreline retreat in the low energy environment (typical of Texas bays) was reduced by providing the cultched shell reef compared to a control site with no reef (Piazza 2005). A study by Meyer et al (1997) showed significant differences in sediment accumulation and marsh edge retreat with cultched and noncultched sites. The sites with oyster reefs added to the lower intertidal fringe of the Spartina alterniflora marshes showed an average addition of 2.9 cm while the unaltered plots showed a retreat average of 1.3 cm over the 1.5 year monitoring http://cdn.coastalcare.org/wp-­‐content/uploads/2012/04/oyster-­‐reef.jpg 15 FIGURE 5: OYSTER REEF ACTING AS A BREAKWATER TO PROTECT COASTAL MARSH IN ALABAMA period (Meyer 1997). Spartina marsh is found along much of the Texas Gulf Coast and is what provides current inundation and surge protection services to many vulnerable coastal areas; the proven ability of oyster reefs to protect and enhance growth of these marshes is a valuable asset of reefs to our coastal communities (Figure 5). Bayous and small river systems that enter Texas bays are also very vulnerable sites affected by storm surges and wave energies; a study conducted on the western end of Mobile Bay in Alabama noted that intertidal oyster reefs constructed in the bayous off the bay reduced wave erosion of the marshes behind the reefs more than nearby natural areas did (Stricklin 2010). This evidence suggests that not only do oyster reefs provide valuable services to coastal shorelines, but they also provide valuable services to more inland waters or those ecosystems which may actually be damaged by levees, jetties, and engineered bulkheads around the boating channels of the bays. The Natural Capital Projects Marine InVEST Team modeled the reduction in wave energy that would result from oyster reef construction at three potential breakwater sites on the northern Gulf Coast using the Coastal Protection Model (Kroeger 2012). It was shown that various reef designs could attenuate a storm wave height by 60-­‐75% and reduce wave energy even more significantly, from 75-­‐99%. Both reefs in the study were shown to reduce the median wave height to below 0.15 m, the threshold for which coastal marshes can survive (Kroeger 2012). T ABLE 5: OYSTER REEF SHORE PROTECTION VALUE SUMMARY TABLE Shore Protection Services Summary Table Source Location Values Scyphers (2011) Mobile Bay, AL 300% increase in shoreline footprint 40% reduction in vegetation retreat Piazza (2005) Louisiana 2-­‐6 cm/month decrease in shoreline retreat rate Meyer (1997) Spartina marsh 4.2 cm/1.5 yr. increase in shore addition Kroeger (2012) Mobile Bay, AL <0.15 m wave height resulting from oyster reef presence Non-­‐Use Value People may benefit from the ƉƌĞƐĞŶĐĞŽĨŽLJƐƚĞƌƌĞĞĨƐŽƌƌĞƐƚŽƌĂƚŝŽŶƉƌŽũĞĐƚƐĞǀĞŶŝĨƚŚĞLJĚŽŶ͛ƚĚŝƌĞĐƚůLJ
consume any of the services mentioned in this paper. These people would derive value from the reefs in the form of existence value and option value, or knowing the reefs exist and being provided the option of using the improved services in the future (Hicks 2004). Hicks estimated this non-­‐use value through surveys to citizens surrounding Chesapeake Bay and concluded that a ten-­‐year, 1000 acre restoration project had a non-­‐use value of at least $114.95 million to the Chesapeake population. This result emphasizes the fact that although it is possible to quantify certain functions and services provided by 16 oyster reefs to the public in terms of both functionality values and economic values, it is difficult to pinpoint exact values due to the complexity and overlapping nature of the services provided by oyster reefs. References Beseres Pollack J, Yoskowitz D, Kim H-­‐C, Montagna PA (2013) Role and Value of Nitrogen Regulation Provided by Oysters (Crassostrea virginica) in the Mission-­‐Aransas Estuary, Texas, USA. PLoS ONE 8(6): e65314. doi:10.1371/journal.pone.0065314 Brumbaugh, R.D., and C.Toropova (2008) Economic valuation of ecosystem services: A new impetus for shellfish restoration? Basins and Coasts News 2 (2):8-­‐15. Coen, L.D. and M. Luckenbach (2000) Developing success criteria and goals for evaluating oyster reef restoration: ecological function or resource exploitation? Ecological Engineering 15: 323-­‐343. Grabowski, J.H., R.D. Brumbaugh, R. Conrad, A.G. Keeler, J. Opaluch, C.H. Peterson, M.F. Piehler, S.P. Powers and A.R. Smyth (2012) Economic valuation of ecosystem services provided by oyster reefs. BioScience. 62:900-­‐909. HĞŶĚĞƌƐŽŶ͕:͕͘ĂŶĚK͛EĞŝů͕>͘:;ϮϬϬϯͿ͞ĐŽŶŽŵŝĐǀĂůƵĞƐĂƐsociated with construction of oyster reefs by ƚŚĞŽƌƉƐŽĨŶŐŝŶĞĞƌƐ͕͟DZZWdĞĐŚŶŝĐĂůEŽƚĞƐŽůůĞĐƚŝŽŶ;ZdE-­‐EMRRP-­‐ER-­‐01), U.S. Army Engineer Research and Development Center, Vicksburg, MS. http://www.wes.army.mil/el/emrrp Hicks, Robert L, Timothy C. Haab, and Douglas Lipton (2004) The Economic Benefits of Oyster Reef Restoration in the Chesapeake Bay. Norfolk, VA: Chesapeake Bay Foundation. Kroeger, T. (2012). Dollars and Sense: Economic Benefits and Impacts from two Oyster Reef Restoration Projects in the Northern Gulf of Mexico. The Nature Conservancy, Arlington, VA. Meyer, D.L., Townsend, E.C., Thayer, G.W (1997) Stabilization and erosion control value of oyster cultch for intertidal marsh. Restoration Ecology 5, 93-­‐99. Millennium Ecosystem Assessment (2005) Ecosystem and human well-­‐being: current state and trends. Washington DC: Island Press. North Carolina Sea Grant (1997) Use of restored oyster reef habitat by economically valuable fishes and crabs in North Carolina: an experimental approach with economic analyses. Project 96FEC-­‐104, Morehead City, NC. Northern Economics, Inc. (2009) Valuation of Ecosystem Services from Shellfish Restoration, Enhancement and Management: A Review of the Literature. Prepared for the Pacific Shellfish Institute 17 Peterson, C.H., Grabowski, J.H., Powers, S.P., (2003) Estimated enhancement of fish production resulting from restoring oyster reef habitat: quantitative valuation. Marine Ecology Progress Series 264, 251-­‐256. Piehler MF, Smyth AR. (2011) Habitat-­‐specific distinctions in estuarine denitrification affect both ecosystem function and services. Ecosphere 2 (art. 12). doi:10.1890/ES10-­‐00082.1 Plunket, J., & La Peyre, M. (2004). Oyster Beds as Fish and Macroinvertebrate Habitat in Barataria Bay, Louisiana. BULLETIN OF MARINE SCIENCE, 77(1), 155-­‐164. (2004, November 5). Retrieved January 1, 2014. Robinson, Lance. (2014) "Oysters in Texas Coastal Waters." TPWD: Short Reports: Oyster Article. Texas Parks and Wildlife Department, n.d. Web. 9 July 2014. <http://www.tpwd.state.tx.us/fishboat/fish/didyouknow/oysterarticle.phtml>. Scyphers SB, Powers SP, Heck KL Jr, Byron D (2011) Oyster Reefs as Natural Breakwaters Mitigate Shoreline Loss and Facilitate Fisheries. PLoS ONE 6(8):e22396. doi:10.1371/journal.pone.0022396 Stricklin, A.G., M.S. Peterson, J.D. Lopez, C.A. May, C.F. Mohrman and M.S. Woodrey.(2010) Do small, patchy, constructed intertidal oyster reefs reduce salt marsh erosion as well as natural reefs? Gulf and Caribbean Research 22:21-­‐27. Stunz, G., Minello, T., & Rozas, L. (2010). Relative value of oyster reef as habitat for estuarine nekton in Galveston Bay, Texas. MARINE ECOLOGY PROGRESS SERIES, 4046, 147-­‐159. (2010, May 10). Retrieved January 1, 2014. Zimmerman, R., T. J. Minello, T. Baumer, and M. Castiglione. (1989) Oyster reef as habitat for estuarine macrofauna. NOAA Tech. Memo. NMFS-­‐SEFC-­‐249. 16 p Zu Ermgassen, P.S.E., M.D. Spalding, B. Blake, L.D. Coen, B. Dumbauld, S. Geiger, J.H. Grabowski, R. Grizzle, M. Luckenbach, K.A. McGraw, B. Rodney, J.L. Ruesink, S.P. Powers, and R.D. Brumbaugh. (2012) Historical ecology with real numbers: past and present extent and biomass of an imperiled estuarine ecosystem. Proceedings of the Royal Society B 279: 3393ʹ3400. Zu Ermgassen, P. S., Spalding, M. D., Grizzle, R. E., & Brumbaugh, R. D. (2013). Quantifying the loss of a marine ecosystem service: filtration by the eastern oyster in US estuaries. Estuaries and coasts, 36(1), 36-­‐43. 18 Coastal Wetlands Coastal wetlands have been degraded greatly in the past decade. A 2013 NOAA report finds that coastal wetlands across the continental U.S. have been lost at an average rate of 80,160 acres per year. Wetland loss in the Gulf of Mexico accounts for nearly 71% of total coastal wetland loss, with an average rate of 57,144 acres per year (Dahl & Stedman, 2013). These figures demonstrate the alarming rate of wetland loss occurring along the coast of Texas, and compel decision-­‐makers to re-­‐examine the value of coastal wetlands. This paper aims to provide a summary of wetland value quantification efforts, and provide an overview of possible conservation actions. Costanza notes his world-­‐wide ecosystem valuation study that tidal wetland values have skyrocketed in ƚŚĞůĂƐƚĚĞĐĂĚĞƐĚƵĞƚŽŶĞǁĞǀŝĚĞŶĐĞƚŚĂƚĚĞŵŽŶƐƚƌĂƚĞƐƚŚĞ͞ƐƚŽƌŵƉƌŽƚĞĐƚŝŽŶ͕ĞƌŽƐŝŽŶĐŽŶƚƌŽů͕ĂŶĚ
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hectare per year, but his 2014 valuation estimates that wetlands are actually worth closer to $194,000 ƉĞƌŚĞĐƚĂƌĞƉĞƌLJĞĂƌ;ŽƐƚĂŶnjĂĞƚĂů͕ϮϬϭϰͿ͘ŽƐƚĂŶnjĂ͛ƐĐĂůĐƵůĂƚŝŽŶƐƌĞůLJŽŶďĞŶĞĨŝƚƚƌĂŶƐĨĞƌŵĞƚŚŽĚƐ͕
pulling from hundreds of studies conducted around the world, and provide a rough estimate of the magnitude of value tidal wetlands provide. In this paper, more attention will be given to studies conducted along the Gulf of Mexico in an attempt to more accurately capture the value of Texas coastal wetlands. Wetland Services Coastal wetlands provide a wide range of services directly in the form of provisioning and cultural services, as well as indirectly through regulating and supporting services. Researchers have identified the following as services provided by coastal wetlands: flood and flow control, storm buffering, erosion control, water quality maintenance, habitat and nursery support for plant and animal species, carbon sequestration, and nutrient cycling as shown in the diagram in Figure 6 (Brander et al, 2006). In addition, coastal wetlands on the Texas coast hold educational, spiritual, historic, and aesthetic value. This paper summarizes relevant studies regarding fish and bird habitat support, carbon sequestration, and F IGURE 6: SERVICES PROVIDED BY A COASTAL WETLAND 19 storm protection. Some services can be traded in markets, and are therefore quantified more easily. However, services that cannot be quantified in market terms pose a challenge. These services are quantiĨŝĞĚƵƐŝŶŐĂůƚĞƌŶĂƚŝǀĞŵĞƚŚŽĚƐ͕ƐƵĐŚĂƐ͞ǁŝůůŝŶŐŶĞƐƐƚŽƉĂLJ͕͟͞ĚĂŵĂŐĞĐŽƐƚƐĂǀŽŝĚĞĚ͕͟ĂŶĚŽƚŚĞƌ
estimation methods. A 1989 study conducted by Costanza attempted to quantify the aggregate value of Louisiana coastal wetlands. This study looked at commercial fishery production, commercial fur trapping production, storm protection, and recreational value using a combination of fish and fur production values, damage costs, and willingness to pay studies. Costanza finds that the aggregate value of Louisiana wetlands is between $2429 and $17000 per acre in 1983 $US (Costanza et al, 1989). In this study, fish and fur support values were easiest to quantify, while recreational value was the most difficult and imprecise. Fishery Support Engle (2011) summarizes literature on Gulf of Mexico coastal wetlands, and explains that many studies have found a link between fish populations and wetland area. Specifically, studies have shown a link between shrimp growth, survival and reproduction, and availability of wetland habitat. Minello et al (2008) explains that shrimp larvae depend on marshland for habitat before growing large enough to move into open water. One metric used to assess wetland fishery support value is population density. This metric is an abstraction of the true value of wetland fishery support, but it is used in many studies because it is straightforward, relatively easy to verify, and applicable to all fishery habitats. Minello & Rozas (2002) developed a model for calculating nekton density in marshes and wetlands that is used as the basis for most subsequent studies on fishery support value. They estimated the density in different regions of marsh by calculating the distance to the nearest vegetated edge. This model assumed that peak nekton density occurs at the edge of a marsh, and decreases with distance from the edge. This model was validated using the Elmgrove Point Marsh in Galveston Bay, and could be applied to most marshes along the Gulf Coast with reasonable accuracy (Minello & Rozas, 2002). Haas et al (2004) found similar results, noting that brown shrimp survive and grow best in regions of high marsh vegetation. Specifically, an abundance of vegetated marsh edges decrease the distance shrimp must travel between open water and vegetation. This reduces movement-­‐related mortality. Minello et al (2008) examined the differences between crustacean growth and survival in wetlands versus the open ocean in Galveston Bay. This paper sampled various locations in Galveston Bay and used size-­‐frequency data for shrimp and blue crabs to find an average mass/individual. Then estimates of crustacean yields in kg/ha were made. The study found that a marsh complex could support 19,382 brown shrimp per hectare, 17,406 white shrimp per hectare, and 16,726 blue crabs per hectare, which represents 3, 2.2 and 4.2 times the standing crop numbers for open water respectively. They estimated that production values for the three species as follows: 128 kg/ha for brown shrimp, 109 kg/ha for white shrimp and 170 kg/ha for blue crabs3. 3
Values are expressed on a lbs. per acre basis in the Wetlands Fishery Support Summary Table 20 Literature from the past decade shows that nekton rely most heavily on marsh edges for nursery habitat. Thus, any attempts to restore or build wetland area should focus on increasing the amount of ǀĞŐĞƚĂƚĞĚĞĚŐĞƐ͘DĂƌƐŚƚĞƌƌĂĐŝŶŐŝƐĂĐŽŵŵŽŶƌĞƐƚŽƌĂƚŝŽŶƉƌĂĐƚŝĐĞ͕ĂŶĚŝŶǀŽůǀĞƐďƵŝůĚŝŶŐ͞ƚĞƌƌĂĐĞƐ͕͟
which are small ridges or levees, in some sort of pattern (usually a checkerboard with open corners) as seen in Figure 7. This creates plenty of marsh edge, which is where the largest density of nekton lives, and the open corners allow for the movement of water and organisms into and out of the marsh. After the ridges are built, they are planted with marsh vegetation, and the space between ridges can be planted with submerged aquatic vegetation. Rozas et al (2005) conducted a study of five different wetland restoration projects along the Galveston coast, and assessed the benefits of each project on fishery habitat. The two terracing projects were the Galveston Island State Park Terracing (GISPT), and the Pierce Marsh Preserve F IGURE 7: TERRACING IN G ALVESTON I SLAND STATE Terracing (PMPT). Both of these projects created P ARK AS SEEN FROM GOOGLE E ARTH terraces in the pattern of a checkerboard in order to maximize edge habitat, and planted smooth cordgrass along the ridges. The Jumble Cove (JC) project took sediment from off-­‐site and used it to create 35 mounds, or islands. The mounds were planted with smooth cordgrass. The two final projects were located at Galveston Island along Interstate Highway 45: I-­‐45 West Marsh (I45WM) and I-­‐45 East Marsh (I45EM). These projects first restored the marsh to its original elevation, and then created narrow channels through the marsh, with smooth cordgrass planted between channels. Each project had different percentages of marsh edge, and the highest was at the terracing projects. Average standing crop for the restored sites was 22,246-­‐30,863 brown shrimp per hectare, 21,773-­‐33,139 white shrimp per hectare, and 17,240-­‐24,927 blue crab per hectare. In contrast, one hectare of open water supports 6431 brown shrimp, 2400 white shrimp, and 7623 blue crab, as reported by Rozas (2005). In some cases the restored marshes increased the standing crop by an order of magnitude. The total construction costs for the different projects ranged from $74,000 to $362,250 depending on the site and the conditions. The GISP terracing project and the PMP terracing project were the most cost-­‐effective in relation to hectare of restored marsh, costing $5,310 per hectare and $8,771 per hectare respectively. The GISP terrace and the PMP terrace also supported the greatest number of nekton per dollar spent on restoration at 9.9 nekton per dollar for the GISP project, and 6.2 nekton per dollar for the PMP project. A later study conducted by Minello et al (2012) looked at nine different marsh restoration projects in Galveston Bay and estimated the value of shrimp production per hectare of restored wetland. Using the annual average shrimp production per hectare, as well as average shrimp prices per kg, Minello found that the annual value of shrimp production of restored marsh ranged from 21 $425-­‐$690/hectare. When compared to the cost of construction of each project, the study found that the terrace projects were the most cost-­‐effective per hectare. Of the five terrace projects evaluated, each took between 18.4 and 20.9 years to recover the cost of construction. Table 6: Wetlands Fishery Support Summary Table Wetlands Fishery Support Study Year Location brown shrimp (ind./ac) white shrimp (ind./ac) blue Production Other values crab Values (ind./ac) (lbs./acre) Project Type Minello 2008 Galveston 47894 43011 41331 695 brown shrimp Rozas 2005 GISPT 65458 61299 44042 9.9 nekton/$ terrace PMPT 67450 63501 44425 592 white shrimp 924 blue crab 6.2 nekton/$ terrace JC 54971 53802 42601 mound I45Wm 76264 81888 61596 channel I45EM 71653 71203 48865 channel Elmgrove 93277 95397 65928 reference Minello 2012 Galveston AVERAGES FROM ALL STUDIES 68138 67157 49827 $171-­‐$279 summary of shrimp 9 projects production/acre After evaluating many of the restoration projects conducted along the Texas Coast, it appears that marsh terracing produces the desired nursery effect while keeping costs at a minimum. This type of marsh restoration has already had substantial success in Galveston Bay, and there is potential for further terracing along the rest of the Gulf Coast. Bird Support Hundreds of species of birds rely on the Texas Gulf Coast for fishing, breeding, and migration. Waterfowl and shorebirds rely on coastal wetlands for feeding grounds, and many northern species winter on the Gulf Coast. The Texas Gulf Coast alone provides habitat for over 100 species of waterbirds. 22 Some studies conducted in the Galveston Bay Estuary have found a link between waterbird populations and marsh area. Gawlik et al (1998) monitored long-­‐term bird populations between 1973-­‐1990, and found that several species declined in population during that time. These species were ones that foraged in vegetated marsh area, and thus the researchers hypothesized that decreasing marsh area in Galveston Bay resulted in declining populations of foraging shorebirds. Freshwater coastal wetlands are particularly vulnerable to loss and degradation. In the past 40 years their area has decreased by 30% along the Texas Coast (Fitzsimmons et al, 2012). Loss of these wetlands could have severe effects on wintering and migratory bird populations, as well as shorebirds that rely on freshwater wetlands. Many efforts have been made to restore and protect coastal wetlands in order to maximize use by a wide variety of birds. Different species of birds prefer different habitat characteristics, so it is imperative that conservation actions provide both vegetated and unvegetated marsh area. Darnell & Smith (2004) studied the use of natural and created salt marshes by bird species at the Aransas National Wildlife Refuge. The researchers looked at shorebirds, perching birds, wading birds, and gulls, and estimated the amount of each type of bird that relied on the salt marsh. Though it was found that the total number of birds present between the natural and created marsh were similar, the makeup of bird type varied significantly. In the created marshes, lack of proper maintenance had allowed vegetation to become overgrown, which decreased the amount of shallow open-­‐water habitat. As a result, fewer waterbirds were observed at these sites, and more perching birds were found. The conclusions of this paper demonstrate that marsh restoration projects must also receive long-­‐term maintenance in order to provide habitat space for a broader range of Texas birds. ŶŽƚŚĞƌƐƚƵĚLJĐŽŶĚƵĐƚĞĚďLJ&ŝƚnjƐŝŵŵŽŶƐĞƚĂů;ϮϬϭϮͿĐŽŵƉĂƌĞĚ͞ŵĂŶĂŐĞĚ͟ŵĂƌƐŚĞƐŝŶ'ĂůǀĞƐƚŽŶĂLJ
ǁŝƚŚ͞ŶŽŶ-­‐ŵĂŶĂŐĞĚ͟ŵĂƌƐŚ͘dŚŝƐƐƚƵĚLJĂƐƐĞƐƐĞĚƚŚĞĞĨĨĞĐƚƐ of salt water intrusion to coastal wetlands by comparing marshes that had levees or barriers installed to protect freshwater wetlands, and marshes ƚŚĂƚŚĂĚŶŽƉƌŽƚĞĐƚŝǀĞƐƚƌƵĐƚƵƌĞƐ͘&ŝƚnjƐŝŵŵŽŶƐĨŽƵŶĚƚŚĂƚŵĂŶĂŐĞĚŵĂƌƐŚĞƐ͞supported more bird species, greater waterbird densities, greater plant diversity, and greater aquatic invertebrate biomass than non-­‐ŵĂŶĂŐĞĚƐŝƚĞƐ͘͟
Since salt intrusion to freshwater coastal wetlands causes a substantial loss of freshwater habitat each year, structural barriers such as the ones studied by Fitzsimmons could be another management action to enhance wetland bird support. http://conservation.audubon.org/sites/default/files/photos/White_Ibis_flock.jpg F IGURE 8: WHITE IBIS FLOCK OVER A COASTAL WETLAND 23 Carbon Sequestration Wetlands are often valued for their carbon sequestration ability because carbon is trapped in living biomass, in non-­‐living biomass, anĚƵŶĚĞƌŶĞĂƚŚŵĂƌƐŚƐĞĚŝŵĞŶƚ͘dŚĞƐĞƋƵĞƐƚĞƌĞĚĐĂƌďŽŶŝƐĐĂůůĞĚ͞ďůƵĞ
ĐĂƌďŽŶ͕͟ĂŶĚĐĂŶďĞƐƚŽƌĞĚŝŶƚŚĞƐŚŽƌƚƚĞƌŵŝŶďŝŽŵĂƐƐ͕ĂŶĚƐƚŽƌĞĚŝŶƚŚĞůŽŶŐƚĞƌŵŝŶƐĞĚŝŵĞŶƚ͘DƵĐŚ
of this carbon is never released back into the atmosphere because marshes vertically accrete, so blue carbon can remain undisturbed while sediments continue to accumulate on top of it (Mcleod et al, 2011). In order to understand the magnitude and rates of carbon sequestration, it is important to quantify the individual carbon sequestration rates of wetland plants, the accumulation rate of carbon in soils and the release rate of carbon dioxide and methane gas from plants and soils during decomposition and anaerobic respiration. Soils sequester the large majority of carbon in wetlands, and some studies estimate soil carbon accounts for 98% of the carbon pool (Engle, 2011). The soil carbon pool is extensive, and can hold an estimated 2300 Pg at a depth of one meter worldwide (Lal, 2002). https://lh6.googleusercontent.com/-­‐6Xr04dBZsRs/UA2L9iiOOkI/AAAAAAAAA0g/oCnmI4ihJUU/s400/pro.jpg F IGURE 9: MARSH TRANSITION TOWARDS UPLAND HABITAT Carbon sequestration in salt marshes varies substantially around the world, and is primarily dependent on sedimentation rates and flooding frequency (Chmura et al, 2003). A global meta-­‐analysis of salt marsh sequestration found that average rates were 210 g CO2 per square meter per year (1874 lbs. CO2/acre/year or 510 lbs. C/acre/year4)(Chmura et al, 2003). However, many subsequent studies have come up with more precise estimates specific to the Gulf of Mexico. Using averages of studies conducted in the Gulf of Mexico, in places like Florida and Louisiana, Engle (2011) estimates that estuarine wetland soils store 275 Mg C per hectare (123 US tons C/acre) in the soil carbon pool. Estuarine wetlands also have high carbon sequestration rates in relation to other wetland types, capturing an average 2.6 Mg C per hectare per year (1.2 tons /acre/year) (Engle, 2011). Another study conducted by Schmidt et al (2014) found that salt marsh along the Georgia coast could store 71 Mg C 4
One pound of carbon dioxide is equivalent to 3.67 pounds of elemental carbon. Mass of elemental carbon sequestered per unit area is another metric often recognized by the scientific community as a standard and what is used throughout this report to describe carbon ecosystem services. 24 ƉĞƌŚĞĐƚĂƌĞ;ϯϭ͘ϲϳƚŽŶƐͬĂĐƌĞͿ͘^ĐŚŵŝĚƚ͛ƐƉĂƉĞƌŶŽƚĞƐƚŚĂƚƚŚŝƐŝƐĂŶƵŶĚĞƌĞƐƚŝŵĂƚĞďĞĐĂƵƐĞŝƚŽŶůLJ
takes into account carbon in the biomass and the first 30cm of soil. Salt marshes can actually store carbon in the soil many meters below the surface. Loomis et al (2010) finds similar results in their carbon analysis of coastal Georgia wetlands, calculating that salt marshes can sequester 162 kg (357 lbs.) C per acre per year. Choi et al (2001) showed that carbon storage in marshes varies depending on elevation. Low marsh can store the most carbon -­‐ an estimated 129 tons C/acre ʹ while high marsh can only store 58 tons C/acre. Choi et al conducted their study along the Gulf Coast of Florida, which has similar conditions as the Texas Gulf Coast. Low elevation soil organic carbon levels on the coast of Florida are reported at 258,732 lbs. C/acre (Choi et al 2001). Figure 9 shows the elevations and distinctions between marsh types. Estuarine wetlands sequester carbon at a rate 10 times higher than other wetlands due to high sedimentation rates, naturally high soil carbon density, and burial of carbon under sediments due to sea level rise. World-­‐wide, estuarine wetlands sequester 11.25 million tons C per year, but have seen a reduction in sequestering capacity of 2.2 million tons C per year. This is due mostly to loss of wetland area (Bridgham et al 2006). Many researchers believe that estuarine wetlands can adapt to sea-­‐level rise by moving landward, growing vertically, and expanding outward into the sea (Choi et al, 2001). , Choi et al (2001) conducted a study to track the inward movement of marsh and observed higher levels of carbon in the soil of low marshes compared to middle and high marshes. As coastal wetlands are pushed back from sea-­‐level rise, the once-­‐inland marshes will make the transition from high marsh to low marsh if landward space is available for them to spread, which would increase carbon sequestration levels (Choi et al 2001). If the rate of landward or new sediment accretion is lower than the rate of sea-­‐level rise, low marsh habitats that are in direct contact with sea-­‐level may slough off, releasing large amounts of stored soil and biomass carbon; it is imperative that as sea levels rise, landward low-­‐elevation land is available for marsh migration inland. The potential for salt marshes to sequester carbon is greatly diminished by anthropogenic activities. Each year human disturbances to coastal wetlands have caused massive amounts of carbon to be released back into the atmosphere. Human disturbances include timber harvesting, clearing, dredging and filling, etc. Mcleod et al (2011) notes that carbon sinks are compromised from these activities in two ways: the carbon stored in plants and other biomass is released back into the air, and carbon stored in sediments is also disturbed and oxidized into the atmosphere. The latter mechanism has the potential to release massive amounts of carbon, since wetland sediments may store several millennia worth of carbon. A study in Malaysia found that a conversion from mangroves to aquaculture ponds resulted in the release of 150 tons C/ha due to the removal of biomass, and another 750 tons C/ha due to the oxidation of mangrove sediment (Mcleod et al, 2011). Clearly, the magnitude of carbon release associated with wetland loss serves as a substantial carbon source that threatens to offset the sequestration abilities of other wetlands. 25 Table 7: Wetlands Carbon Services Summary Table Wetland Carbon Sequestration Study Year Location Carbon Sequestration Storage Rate (tons C/acre) (lbs./acre/yr.) Notes Engle 2011 Gulf of Mexico 114 2319 Meta-­‐analysis of wetland studies conducted along the Gulf Coast Chmura et al 2003 World-­‐wide 174 508 Carbon Sequestration rates are global averages Choi et al 2001 Florida (low marsh) 129 8465 Closest to open water Choi et al 2001 Florida (middle marsh) 67 5307 Choi et al 2001 Florida (high marsh) 58 2168 Farthest landward Schmidt et al 2014 Georgia 31 Only measured the biomass of first 30 cm of soil Loomis et al 2010 Georgia 357 Conservative estimate AVERAGES FROM ALL STUDIES 96 3187 Storm Protection Coastal wetlands are highly regarded for their ability to provide a buffer between hurricane storm surge and coastal infrastructure. In particular, the roughness of marsh vegetation, as well as shallow water depths associated with coastal wetlands, are able to dissipate wave energy and attenuate storm surge. Researchers took note of this phenomenon starting in 1963, when the Army Corps of Engineers published a report documenting surge attenuation in wetland areas (Engle, 2011). Since then, researchers have aimed to quantify attenuation potential, and understand the relationship between coastal vegetation and surge levels. Wetland vegetation is able to attenuate surges by exerting a drag 26 force on the flowing water, reducing turbulence, and slowing water velocity. Studies have shown that this effect happens whether vegetation is partially or deeply submerged. In addition, wetland vegetation indirectly reduces surge levels over several decades by building up peat and changing the shore bathymetry (Gedan et al, 1997). Researchers have used both documented hurricane data, as well as validated numerical models to calculate surge attenuation. Along the Gulf Coast, wetlands have been shown to attenuate surge at a rate of up to 1 m surge reduction per 7.4 km wetland traverse (Engle, 2011). Wamsley et al (2010) ran four simulated storms across two wetland swatches in Louisiana and tracked attenuation rates. After running each simulated storm, researchers found that surge was attenuated between 1m per 25 km of wetlands marsh and 1 m per 6 km of wetland marsh. This large range of rates demonstrates that surge attenuation is greatly variable, even for wetlands with similar characteristics. Specific bathymetric features, as well as storm intensity, are features that largely affect attenuation rates. The researchers conclude that though wetlands attenuate surge significantly, the traditional linear expression of this phenomenon is imprecise and unreliable for future predictions. Perhaps a more precise metric for calculating surge attenuation is a wetland-­‐to-­‐water ratio. By calculating the relative amount of vegetation and the wetland roughness, researchers can predict a ǁĞƚůĂŶĚ͛ƐƉŽƚĞŶƚŝĂůƚŽĂƚƚĞŶƵĂƚĞƐƵƌŐĞ͘ĂƌďŝĞƌĞƚĂů;ϮϬϭϯͿĐŽŵƉĂƌĞĚƐƚŽƌŵƐƵƌŐĞĞůĞǀĂƚŝŽŶƐĨŽƌĂŶĂƌĞĂ
of open water versus an area of highly vegetated marsh. Using an ADCIRC hydrodynamic model, four different hypothetical storms were run along the Caernarvon Basin in Louisiana. It was found that a 1% increase in wetland-­‐to-­‐water ratios decreased surge by 8.4-­‐11.2%, and a 1% increase in wetland ͞ƌŽƵŐŚŶĞƐƐ͟ĚĞĐƌĞĂƐĞĚƐƵƌŐĞďLJϭϱ͘ϰ-­‐28.1%. Though this method of quantification is not as widely used, ŝƚŵŝŐŚƚďĞŵŽƌĞĂƉƉůŝĐĂďůĞƐŝŶĐĞŝƚŵĞĂƐƵƌĞƐĂǁĞƚůĂŶĚ͛ƐƉŽƚĞŶƚŝĂůƚŽĂƚƚĞŶƵĂƚĞƐƵƌŐĞƌĂƚŚĞƌƚŚĂŶ
quantified surge reductions, which are highly variable due to the multiple factors that control surge levels. Efforts to quantify wetland protection value in dollars have also been undertaken. Usually, wetland values are calculated using damage costs, or damage costs avoided. Barbier et al (1997) boldly calculated that a loss of coastal wetlands in Louisiana of one mile inland would result in increased hurricane damage costs of an estimated $5.75 million annually. A more realistic estimate by Barbier et al (1997) calculates that each acre of lost wetland corresponds to $128 of increased damage annually. Costanza et al (2008) conducted a nation-­‐wide study of coastal wetlands, and found that a loss of one hectare of wetland area resulted in damage increases of an average of $33000 per year. By mapping hurricane probabilities across the country, it was found that the average value of coastal wetland was $8240/ha per year. Along the Texas coast, Costanza estimated a higher value of $12365/ha per year. Though these numbers would not correspond to actual market values, they serve as indicators of the importance of wetland integrity along the Gulf Coast. Coastal and Estuarine wetlands of the Texas Gulf Coast provide a diverse range of benefits, from regulating gas exchange, to protecting coastal communities from hurricane surge, to supporting the rich biodiversity of coastal fish and birds. Past Quantification methods have demonstrated that these benefits are substantial, but often go unnoticed by decision-­‐makers in society. In order to assess the 27 true value of coastal wetlands and marshes, and in order to understand the full consequences of wetland degradation, these metrics should be taken into consideration. Though additional research efforts should aim to refine quantification methods, the metrics presented in this paper serve as a starting range of values to work from. References Barbier, E. B., Acreman, M. C., & Knowler, D. (1997). Economic valuation of wetlands: a guide for policy makers and planners. Ramsar Convention Bureau, Gland, Switzerland. Barbier, E. B., Georgiou, I. Y., Enchelmeyer, B., & Reed, D. J. (2013). The Value of Wetlands in Protecting Southeast Louisiana from Hurricane Storm Surges. PLoS ONE, Vol 3, No. 3. Brander, L. M., Florax, R. J. G. M., & Vermaat, J. E. (2006). The Empirics of Wetland Valuation: A Comprehensive Summary and a Meta-­‐Analysis of the Literature. Environmental & Resource Economics, Vol 33, 223-­‐250. Bridgham, S. D., Megonigal, P. J., Keller, J. K., Bliss, N. B., & Trettin, C. (2006). The Carbon Balance of North American Wetlands. Wetlands Vol. 26, No. 4, 889-­‐916. Chmura, G. L., Anisfield, S. C., Calhoun, D. R. & Lynch, J. C. (2003). Global carbon sequestration in tidal, saline wetland soils. Global Biogeochemical Cycles, Vol 17, No. 4. ŽƐƚĂŶnjĂ͕Z͕͘Ě͛ƌŐĞ͕Z͕͘ĚĞ'ƌŽŽƚ͕Z͕͘&ĂƌďĞƌ͕^͕͘'ƌĂƐƐŽ͕D͕͘,ĂŶŶŽŶ͕͕͘ǀĂŶĚĞŶĞůƚ͕D͘;ϭϵϵϳͿ͘dŚĞ
ǀĂůƵĞŽĨƚŚĞǁŽƌůĚ͛ƐĞĐŽƐLJƐƚĞŵƐĞƌǀŝĐĞƐĂŶĚŶĂƚƵƌĂůĐĂƉŝƚĂů͘Nature, Vol 387, 253-­‐260. Costanza, R., de Groot, R., Sutton, P., vĂŶĚĞƌWůŽĞŐ͕^͕͘ŶĚĞƌƐŽŶ͕^͘:͕͘<ƵďŝƐnjĞǁƐŬŝ͕/͕͙͘dƵƌŶĞƌ͕Z͘<͘
(2014). Changes in the global value of ecosystem services. Global Environmental Change, Vol 26, 152-­‐158. Costanza, R., Perez-­‐Maqueo, O., Martinez, M. L., Sutton, P., Anderson, S. J., & Mulder, K. (2008). The Value of Coastal Wetlands for Hurricane Protection. Ambio, Vol 37, No. 4, 241-­‐248. Dahl, T. E., & Stedman, S. M. (2013). Status and trends of wetlands in the coastal watersheds of the Conterminous United States 2004 to 2009. U.S. Department of the Interior, Fish and Wildlife Service and National Oceanic and Atmospheric Administration, National Marine Fisheries Service, 46 Darnell, T. M., & Smith, E. H. (2004). Avian Use of Natural and Created Salt Marsh in Texas, USA. Waterbirds, Vol 27, No. 3, 355-­‐361. Engle, V. D. (2011). Estimating the Provision of Ecosystem Services by Gulf of Mexico Coastal Wetlands. US Government. 28 Fitzsimmons, O. N., Ballard, B. M., Merendino, M. T., Baldassarre, G. A., Hartke, K. M. (2012). Implications of Coastal Wetland Management to Nonbreeding Waterbirds in Texas. Wetlands, Vol 32, 1057-­‐1066. Gawlik, D. E., Slack, R. D., Thomas, J. A., & Harpole, D. N. (1998). Long-­‐term Trends in Population and Community Measures of Colonial-­‐nesting Waterbirds in Galveston Bay Estuary. Waterbirds, Vol 21, No. 2, 143-­‐151. Gedan, K. B., Kirwan, M. L., Wolanski, E., Barbier, E. B., & Silliman, B. R. (2010). The present and future role of coastal wetland vegetation in protecting shorelines: answering recent challenges to the paradigm. Climatic Change, Vol 106, 7-­‐29. Haas, H. L., Rose, K. A., Fry, B., Minello, & T. J., Rozas, L. P. (2004). Brown Shrimp on the Edge: Linking Habitat to Survival Using an Individual-­‐Based Simulation Model. Ecological Applications, Vol 14, 1232-­‐1247. Lal, R. (2002). Soil erosion and the global carbon budget. Environment International, Vol 29, 437-­‐450. Loomis, M. J., & Craft, C. B. (2010). Carbon Sequestration and Nutrient (Nitrogen, Phosphorus) Accumulation in River-­‐Dominated Tidal Marshes, Georgia, USA. Spo; Science Society of America Journal, Vol 74, No 3, 1028-­‐1036. DĐůĞŽĚ͕͕͘ŚŵƵƌĂ͕'͘>͕͘ŽƵŝůůŽŶ͕^͕͘^Ăůŵ͕Z͕͘ũŽƌŬ͕D͕͘ƵĂƌƚĞ͕͘D͕͙͘^ŝůůŝŵĂŶ͕͘Z͘;ϮϬϭϭͿ͘
blueprint for blue carbon: toward an understanding of the role of vegetated coastal habitat in sequestering CO2. Front Ecol Environ, Vol 9, 552-­‐560. Minello, T. J. Rozas, L. P., Caldwell, & P. A., Liese, C. (2012). A Comparison of Salt Marsh Construction Costs with the Value of Exported Shrimp Production. Wetlands, Vol 32, 791-­‐799. Minello, T. J., Matthews, G. A., Caldwell, P. A., & Rozas, L. P. (2008). Population and Production Estimates for Decapod Crustaceans in Wetlands of Galveston Bay, Texas. Transactions of the American Fisheries Society, Vol 137, 129-­‐146. Minello, T.J., Able, K. W., Weinstein, M. P., & Hays, C. G. (2003). Salt marshes as nurseries for nekton: testing hypothesis on density, growth, and survival through meta-­‐analysis. Marine Ecology Progress Series, Vol 246, 39-­‐59. Rozas, L. P., Caldwell, P., & Minello, T. J. (2005). The Fishery Value of Salt Marsh Restoration Projects. Journal of Coastal Research, Special Issue No. 40, 37-­‐50. Rozas, L. P., & Minello, T. J. (2001). Marsh Terracing as a Wetland Restoration Tool for Creating Fishery Habitat. The Society of Wetland Scientists, Vol 21, 327-­‐341. 29 Schmidt, J. P., Moore, R., & Alber, M. (2014). Integrating ecosystem services and local government finances into land use planning: A case study from coastal Georgia. Landscape and Urban Planning, Vol 122, 56-­‐67. Wamsley, T. V., Cialone, M. A., Smith, J. M., Atkinson, J. H., & Rosati, J. D. (2009). The potential of wetlands in reducing storm surge. Ocean Engineering, Vol 37, No. 1, 59-­‐68. 30 Coastal Bottomland Hardwood Forests The bottomland hardwood forests of the Texas coast represent a unique forest system that provides valuable ecosystem services to the surrounding rural regions and expanding urban population. The freshwater streams and rivers that flow F IGURE 10: C OLUMBIA BOTTOMLANDS CONSERVATION AREA through this bottomland forest and the gulf waters and estuaries that these freshwater systems empty to also heavily depend on the ecosystem functions that are provided by the densely wooded areas of coastal hardwood forests and bottomland wetland forests (Figure 10). This area of forested coastal lands is known as the Columbia Bottomlands and has withstood decades of development, timber harvest, and agricultural clearing to endure as the largest remnant expanse of forested land along the Gulf Coast (U.S. Fish and Wildlife Service 1997). However, this area is coming under increasing threat along with its valuable ecosystem services due to continued fragmentation, agricultural development, urban expansion along the coast, and encroaching sprawl from surrounding communities emerging from the greater ,ŽƵƐƚŽŶĂƌĞĂ͛ƐĐŽƌĞ͘ĞĐĂƵƐĞŽĨƚŚŝƐŝŵŵŝŶĞŶƚ
threat to the last remaining expanses of Texas coastal bottomland forest, it is vital that we understand and recognize the ecosystem services provided to us by these forested lands so they may be valued and preserved for continued enjoyment and utilization. Source: The Nature Conservancy, Land Stewardship in the Columbia Bottomlands Columbia Bottomlands of Texas: A Guide for Landowners. The ͞Columbia Bottomlands͟ is the commonly referred to name given to the vast area of riparian and bottomland hardwood coastal forests along the central Texas coast after Stephen F. Austin first explored the area in 1823 and settled ŝŶĂƚŽǁŶǁŝƚŚŝŶƚŚĞďŽƚƚŽŵůĂŶĚƐŚĞĐĂůůĞĚ͞ŽůƵŵďŝĂ͕͟ŶŽǁĐĂůůĞĚtĞƐƚ
ŽůƵŵďŝĂ͕ǁŚŝĐŚĂĐƚĞĚĂƐŽƵƌƐƚĂƚĞ͛ƐĨŝƌƐƚĐĂƉŝƚĂů͖ƉŽƌƚŝŽŶƐŽĨƚŚĞďŽƚƚŽŵůĂŶĚƐĂƌĞĂůƐŽĐĂůůĞĚ͞ƵƐƚŝŶ͛Ɛ
Woods͟ by the U.S. Fish and Wildlife Service when referring to specific units of the Texas Mid-­‐coast National Wildlife Refuge Complex (Dawson 2004, U.S. Fish and Wildlife Service 1997). Within the Columbia Bottomlands Conservation Area, as designated by The Nature Conservancy, lie several state and national wildlife parks and refuges such as Stephen F. Austin State Park, Brazos Bend State Park, Big Boggy National Wildlife Refuge, Justin Hurst Wildlife Management Area, Nannie M. Stringfellow Wildlife 31 Management Area, San Bernard National Wildlife Refuge, Brazoria National Wildlife Refuge, Peach Point Wildlife Management Area, and some others. This area has been recognized by the state of Texas, federal agencies, and world-­‐renowned conservation agencies as an important natural asset to the Texas coast that must be cherished and protected. The bottomland hardwood forests of the east Texas coast once occupied an area of more than 700,000 acres, stretching across the floodplains of the San Bernard, Brazos and ŽůŽƌĂĚŽZŝǀĞƌ͛ƐĨůĂƚĂůůƵǀŝĂůƐŽŝůƐ
down to the end of the watershed at the Gulf of Mexico shoreline. These floodplains expand across four counties of Texas: Brazoria, Matagorda, Fort Bend, and Wharton. This expanse of floodplains and forests has created a completely unique ecosystem over centuries of fresh water flow from upstream rivers, flooding and forest growth succession, different even from other bottomland hardwood forests along the northern Gulf of Mexico coastal states. With flooded sloughs that cut through the land as remnants of changing river and bayou channels, shallow remnant banks of historical streams, and three dominant ƌŝǀĞƌƐLJƐƚĞŵƐƚŚĂƚĞĂƐŝůLJĨůŽŽĚŝŶƚŽŽŶĞĂŶŽƚŚĞƌ͛ƐƚƌŝďƵƚĂƌŝĞƐĂŶĚǁĂƚĞƌƐŚĞĚƐŽǀĞƌƐŚĂůůŽǁďĂŶŬƐĂŶĚ
across flat, alluvial clay wetland areas, the coastal bottomland forests and wetlands of Texas are a diverse and consistently dynamic ecosystem (Dawson 2004). There are many terms used in literature to describe the Columbia Bottomlands of Texas: bottomland hardwood forests, riparian wetland forests, fluvial woodlands, riparian hardwood forest, coastal wetland forests, and other combinations of terms; but they all refer to the same unique and diverse assemblage of rivers, tributaries, bayous, sloughs, wetlands, banks, floodplains, and diverse understory plants and overstory hardwood forest. Ecosystem Services of a Bottomland Hardwood Forest Costanza et al. (1997) estimated the global value of ecosystem services and determined that swamps and floodplains had the second highest economic value at $7,927/acre/ year, second only to coastal ĞƐƚƵĂƌŝĞƐĂƚΨϵ͕ϮϰϴͬĂĐƌĞͬLJĞĂƌ͘ůƚŚŽƵŐŚŽƐƚĂŶnjĂ͛ƐďƌŽĂĚĐůĂƐƐŝĨŝĐĂƚŝŽŶŽĨ͞ƐǁĂŵƉƐĂŶĚĨůŽŽĚƉůĂŝŶƐ͟
encompasses ecosystems other than bottomland hardwood forests, the bottomlands are nonetheless considered economically and ecologically valuable. The Millennium Ecosystem Assessment of 2005 classifies ecosystem services into four categories: provisioning, regulating, supporting, and cultural. The provisioning services of a forest, specifically a hardwood forest such as the Columbia Bottomlands, include the harvest of timber for logging or chipping. Regulating services of a bottomland hardwood forest include groundwater supply from the inundation of floodwaters over the natural landscape and the flood and hurricane protection this inundation capacity provides, as well as Soils
nutrient control, carbon cycling and sequestration, and F IGURE 11: T YPICAL FUNCTIONS OF A BOTTOMLAND WETLAND FOREST (SUN 2002) 32 ŽƚŚĞƌƉŽůůƵƚĂŶƚƌĞŵŽǀĂůƉƌŽǀŝĚĞĚďLJǁŽŽĚLJƉůĂŶƚƵƉƚĂŬĞ͕ƌŽŽƚĨŝůƚƌĂƚŝŽŶ͕ĂŶĚƚŚĞĚĞŶƐĞĐůĂLJƐŽŝů͛ƐĂďŝůŝƚLJ
to adhere particles. Supporting functions of the forest tend to be the most publically recognized and include the valuable habitat provided by the bottomlands neo-­‐tropical migrant birds, waterfowl, and other wildlife such as white-­‐tailed deer, and many reptiles and amphibians. This valuable and unique habitat provided by the bottomlands to hundreds of migrant bird species also supports the cultural and recreational services provided by the bottomlands. These patches of coastal forest provide a world-­‐
renowned bird-­‐watching site during the fall and spring migration as well as year-­‐round hiking, paddling, and scenic tour opportunities for local residents and outdoor travelers. The Columbia Bottomlands provide the setting for many parks, paddling trails, and National Wildlife Refuges that support the local tourism economy and provide cultural benefits to anyone who visits these sites. There are countless reasons why the Columbia Bottomlands hardwood forests are valuable to the Texas ĐŽĂƐƚ͛ƐƌĞƐŝĚĞŶƚƐ͕ĞĐŽƐLJƐƚĞŵƐ͕ǁŝůĚůŝĨĞ͕ĂŶĚƐƵƌƌŽƵŶĚŝŶŐĐŽŵŵƵŶŝƚŝĞƐ͘dŚĞďĂƐŝĐĞĐŽůŽŐŝĐĂůĨƵŶĐƚŝŽŶƐŽĨĂ
bottomland forest that we benefit from are displayed in Figure 11 (Sun 2002). This review will focus on a few of the most studied and quantifiable services provided by this bottomland riparian forest: inundation capacity of the undeveloped land for surge and flood control, reduction of nitrogen and other nutrients from agricultural and urban runoff into freshwater systems, carbon sequestration rates from growth of hardwood species and carbon sink capacity, and the value of habitat provided to migratory bird species. Inundation Capacity The Columbia Bottomlands of the Texas coast were once a solid and extensive forest and riverine floodplain ecosystem. Over the past several decades, drainage for development, agricultural use, and continuous fragmentation have reduced the extent of the bottomlands from 700,000 acres to a patchwork network of about 150,000 acres, less than 25% of what was once there (U.S. Fish and Wildlife Service 1997, 2013). A study for the US Department of Agriculture on the soil characterization of the Columbia Bottomlands generally concluded that the conditions of the soil and hydrology of the area make it a very difficult region to identify and delineate jurisdictional wetlands, mainly because areas that are jurisdictional can be hard to identify based on the presence of hydric soils and wetland hydrology alone (Miller and Bragg 2007). The presence of hyrdophytic vegetation is evident beyond the recognizable wetland boundaries, indicating that areas not possessing obvious wetland properties could still be considered wetland areas. The main source of wetland hydrology that is evident in the bottomlands is ponding of low lying areas after heavy rain (Miller and Bragg 2007). The study did observe indicators of wetland hydrology at various sites within the bottomlands including water marks on trees, drifted leaf litter, and evidence of ponding such as matted leaves and sparse ground cover in depressed areas (Miller and Braggs 2007). The study concluded that areas within the bottomland hardwood forest that are subject to frequent ponding or flooding during the growing season lack traditional hydric soils and wetland properties, but still may be considered jurisdictional wetlands. The unique soil properties and vegetation characteristics of the Columbia Bottomlands present an issue when areas of the forest are faced with development if not jurisdictionally classified as a wetland and thus protected under wetland conservation regulations. During dry periods, when wetland characteristics are not as prominent, areas that may have inundation capability or exhibit wetland hydrology conditions during wet periods can be easily cleared, filled, covered, and developed-­‐ a process that has gradually fragmented and decreased the total area of the bottomland wetland forest. Hurricane Ike struck the Texas coast in 2008 as a strong Category 2 hurricane with maximum winds of 110 mph at landfall. Storm surge was pushed as far as 30 miles inland in parts of eastern Texas and southwestern Louisiana, and in parts of Galveston Bay, storm surge reached heights of nearly 15 feet and reported 20 ft. high water line due to wave action along Bolivar peninsula (Berg 2009). Ike brought 33 into focus the concern of coastal communities and industries about the risk of flooding and damage from storm surge due to tropical storms or hurricanes in the Gulf of Mexico. Although this particular ŚƵƌƌŝĐĂŶĞ͛ƐƉĂƚŚƚŽŽŬŝƚŵŽƌĞŶŽƌƚŚƚŚĂŶƚŚĞŵŽƐƚŚĞĂǀŝůLJƉŽƉƵůĂƚĞĚĂŶĚŝŶĚƵƐƚƌŝĂůŝnjĞĚƉŽƌƚŝŽŶƐŽĨƚŚĞ
coast, it nevertheless shined a spotlight on the damaging effect of storm surge flooding and has made researchers and city planners more aware of the need for natural areas that can accommodate large levels of inundation. Hurricane Katrina created a reported 27.8 ft. storm surge in Pass Christian, Mississippi when it made landfall in 2005 and the record for the Texas coast is a reported 22.8 ft. in Port Lavaca from Hurricane Carla in 1961 (Masters 2014). These record levels further indicate the risk level the entire Texas coast faces every hurricane season. The previously-­‐mentioned USDA soil study of the bottomland hardwood region of the Texas coast indicates that the hydrology and soil profile of the area is conducive to inundation and water retention (Miller and Bragg 2007). The soils in the area are classified as Vertisols and contain high ratios of clay particles; the Vertisol soil cracks deeply during dry periods and then absorbs water quickly through the cracked pores and swelling into the clay, filling the cracks and sealing the soil surface which allows for ponding of fresh surface water (Miller and Bragg 2007). As shown in Figure 12, roughly half of the bottomland forests of the Columbia Bottomlands occur below the 25 foot elevation contour; in a severe storm surge flooding event, low-­‐elevation undeveloped bottomland forest habitat has the ability to buffer and retain high levels of storm surge that would otherwise flood property and industrial Sources: elevation contour (H-­‐GAC) bottomland area (Houston Wilderness) complexes if developed, F IGURE 12: ABOUT HALF OF THE AREA OF THE BOTTOMLANDS potentially reducing damage FALLS BELOW 25 FOOT ELEVATION to surrounding and upland communities significantly. Low-­‐elevation forested land is also found along major rivers, tributaries, and in their floodplains, demonstrating that the Columbia Bottomlands is an important resource for absorbing flood waters and overbank flow of those river systems during intense rainfall events and generally wet seasons. Land use change and land cover conversion to impervious cover, or even decreased area of woody vegetation for agricultural or grazing purposes, can have cascading effects on the hydrologic regime of ecosystems down-­‐watershed of the changes, including alteration of vegetation structure, habitat suitability, atmospheric consequences, and nutrient control (Faulkner 2004). As Houston and surrounding towns continue to grow, develop and convert natural prairie land and floodplains into impervious cover with buildings and pavement, flooding and intense surface-­‐water flows from heavy rainfall events will become more frequent and damaging to communities and ecosystems that lie downstream. Impervious cover in the upper watersheds will force water that is naturally attenuated and absorbed by plants and soils, or even delayed by being held in floodplains or low-­‐lying wetlands, to 34 travel downstream through the watershed quicker and more intensely as shown in the graph in Figure 13. The presence of bottomland hardwood forests, shallow sloughs and floodplain wetlands that are found throughout the Columbia Bottomlands can attenuate flow coming from up-­‐watershed, store floodwaters, recharge groundwater as the water slowly meanders through, and adsorb large quantities into its clay soils and hardwood trees. Although the bottomlands have the capability to handle inundation for long periods of time to provide us with the protection from flooding and storm surge, increasing the frequency of flooding or even decreasing the amount of freshwater flow into the system from diversion to irrigation or newly-­‐formed urban communities over time can drastically alter other ecosystem services provided by these bottomland hardwood forests. Additionally, a direct disturbance to the bottomland habitat along the coast has an equivalent effect on the hydrology and water quality of down-­‐stream ecosystems such as coastal wetlands and estuaries (Faulkner 2004). Texas coastal populations are growing at a rate of 22% per year, increasing the threat of fragmentation of the bottomland forests (Barrow 2005). The ability to store freshwater and slowly release it into the gulf also maintains delicately balanced estuaries and salt water marshes and wetlands that provide additional hurricane protection benefits. These coastal ecosystems have all developed synchronously over time, relying on a delicate balance that can be thrown off with the slightest disturbance to any one part. Ecosystem services provided to us by coastal ecosystems such as the Columbia Bottomlands are tremendously undervalued and under-­‐
recognized. The equilibrium between these ecosystems and anthropological development is becoming Source: Meiyin Wu, 2013 increasingly unbalanced and http://www.nj.com/morris/index.ssf/2013/04/speaker_urban_development_and.html must be maintained if we hope to continue to utilize F IGURE 13: T HE CHANGE IN PEAK FLOWS WHEN A FORESTED WATERSHED IS the free, natural services that CONVERTED TO URBAN LAND are provided to us by coastal ecosystems. 35 Nutrient and Pollutant Reduction from Runoff Overenrichment of nutrients from human pollutant and agricultural sources in to coastal ecosystems is one of the major stresses of the delicate balance coastal ecosystems must sustain to continue to prosper and provide us with continuing services. Nutrients such as nitrogen and phosphorous, in the form of nitrate and phosphate, are essential for all living things to survive in any ecosystem, but excessive nitrogen and phosphorus degrade surface water and coastal systems water quality. Excess nutrients cause increased algal population growth and increases the amount of organic carbon within a water body which depletes oxygen level; this process is known as eutrophication. Depleted oxygen levels in ƐŽŵĞĂƌĞĂƐŽĨƚŚĞ'ƵůĨŽĨDĞdžŝĐŽŚĂǀĞůĞĚƚŽǁŚĂƚŵĂŶLJƉĞŽƉůĞŬŶŽǁĂƐ͞ĚĞĂĚnjŽŶĞƐ͕͟ŽƌĂƌeas of hypoxia that contain less than 2 mg/L of oxygen, which most aquatic species cannot survive in (Mitsch 2001). Hypoxia zones in the Gulf of Mexico threated commercial fisheries and the economies of coastal states that largely depend on the fishing industry for support. The issue of eutrophication of gulf waters stemming from excess nutrient loads in the Mississippi River traveling into the gulf has spurred a large concern for other gulf states such as Texas. The anthropogenic sources of nutrients that are causing hypoxia to occur in large areas off the coast of Louisiana are similarly present along much of the coast of Texas, and therefore pose a similar threat to our coastal waters and fisheries. Figure 14 shows the large hypoxic zone off the coast of Louisiana that has been a cause of concern for many years as well as a vulnerable area on the coast of Texas, near the outlet of the Colorado and Brazos River systems that flow through the Columbia Bottomlands. Nitrate and phosphate are commonly used in fertilizer and are also a result of high volumes of animal waste and decomposing plant matter that is produced through many agronomic processes. Nitrate and excess particulate phosphorus is quickly accumulated in agricultural soils where fertilizer is commonly used and livestock is raised and urban communities that produce large volumes of sewage and use fertilizer on residential and commercial lawns. It leaches into Source:http://water.epa.gov/type/watersheds/named/msbasin/images/hypoxiawatch_2013.jpg F IGURE 14: HYPOXIC ZONES SHOWN IN RED ALONG THE COASTS OF LOUISIANA AND PARTS OF TEXAS 36 groundwater and runs off into surface water during rain events, eventually being washed through the watershed and accumulating in surface water sources such as streams, rivers, and eventually the ocean. Even with higher levels of anthropogenic sources of nutrients, surface water, groundwater, and coastal waters can be protected from excess nutrient accumulation through natural denitrification and nutrient uptake processes provided by riverine forests, bottomland wetlands, and coastal marshes. In a report to the governor of Louisiana, a group of scientists analyzed the ecosystem services provided by the forested wetlands along the Louisiana coast and concluded that nitrate removal from non-­‐point pollution sources by the forested wetland systems was a significant service provided by this particular ecosystem (Coastal Wetland Forest Conservation and Use Science Working Group 2005). However, a detrimental patter has emerged in that as development continues at a high rate, we are simultaneously creating more nutrient pollution sources as we degrade and destroy the ecosystems that provided us with natural protection against eutrophication of our waters. The shallow floodplains of the Columbia Bottomlands combined with the Vertisol soils that promote frequent ponding and common presence of low-­‐lying areas and remnant stream sloughs do create the proper environment for prolonged inundation and anaerobic soil conditions (shown in Figure 15) that would significantly contribute to denitrification of surface water that flows through the system after heavy runoff events and of the shallow groundwater extending from the coupled river systems that carry pollutants downstream. The anaerobic conditions of wetland soils create the opportunity for microbial populations to reduce nitrate (NO3-­‐) to nitrogen gas (N2) or nitrous oxide (N2O), which removes the nitrogen from the system as it is expelled as a gas from the soil and ponded water (Coastal Wetland Forest Conservation and Use Science Working Group 2005). Organic nitrogen that is not subject to denitrification is also assimilated into plant tissue at high rates through the extensive root system into a Source: http://steveandjudystravelblog.blogspot.com/2011/03/brazos-­‐bend-­‐state-­‐park.html F IGURE 15: I NUNDATED AREA OF B RAZOS BEND STATE P ARK 37 growing forest wetland. Phosphorous, which does not have a gaseous form, and other heavy metals that can be found in non-­‐point source pollution runoff or urban rainfall runoff become attached to the fine particles and sediment within the forested wetlands and are retained through the natural sedimentation process that the dense vegetation of the bottomland forests promote. Through these processes, the bottomland wetland and riparian forests of the mid-­‐Texas coast are connected to surrounding upland agricultural ecosystems and urban communities that act as pollution source to effectively buffer the flow of harmful compounds and pollutants to the coastal ecosystems further downstream (Coastal Wetland Forest Conservation and Use Science Working Group 2005). While natural forested wetlands generally have a high capability for denitrification, it has been shown that some restored wetlands can have inadequate hydrology and limited available carbon and microbial populations compared to natural systems which results in much lower denitrification rates (Hunter and Faulkner 2001). A study of the bottomland hardwood wetlands in Louisiana reported a denitrification rate of a natural bottomland hardwood forest of 8.2 grams NO3-­‐ -­‐ N per kilogram of soil per year, which was significantly higher than the rates reported from restored wetland habitat with and without reestablished hydrology, 5.7 g NO3-­‐ -­‐ N and 1.4 g NO3-­‐ -­‐ N respectively (Hunter and Faulkner 2001). This data stresses the importance of preserving the limited area natural bottomland habitat still existent along the coast, and the importance that complete hydrologic regime in the bottomland plays for water quality functions. Because of the infrequency of complete inundation of the bottomland forests of the area, one might ĐŽŶƐŝĚĞƌƉŽƌƚŝŽŶƐŽĨƚŚĞŽůƵŵďŝĂŽƚƚŽŵůĂŶĚƐ͕ĞƐƉĞĐŝĂůůLJƚŚŽƐĞĞdžƚĞŶĚŝŶŐĂůŽŶŐƚŚĞŵĂũŽƌƌŝǀĞƌ͛Ɛ
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Chesapeake Bay watersheds looked at several different studies conducted over the past few decades. Analysis of one study out of Georgia presented estimates of denitrification rates from riparian forests that ranged from as low as 1.2 lbs. N/acre/year from an area adjacent to an abandoned field that did not receive any fertilizer treatment, 27.6 lbs. N/acre/year for the entire riparian zone to a maximum rate of 263 lbs. N/acre/year from conditions of high nitrate and carbon loads from pig farm nearby (Hendrickson 1981, Lowrance 1984, Lowrance et al 1997). These rates indicate that systems exposed to higher concentrations of nutrient pollution will exhibit higher denitrification and retention rates than those subjected to lower loads. The report also displayed rates of woody uptake of nutrients in coastal plain riparian forests from several studies. Those rates ranged from 10-­‐19 lbs. N/acre/year in woody storage and 68-­‐74 lbs. N/acre/year in total uptake for nitrogen and 2.6-­‐4.4 lbs. phosphorus per acre per year in woody storage and 6.7-­‐8.9 lbs. P/acre/year in total uptake (Lowrance et al. 1997). Lowrance et al concluded through their evaluation that an established riparian forest buffer system can remove 85%-­‐
90% of nitrate input in shallow groundwater which was 17-­‐35 lbs. of Nitrate per acre per year for the systems they studied; total nitrogen retention in established buffer systems was 67%-­‐89% or 23-­‐66 lbs. N/acre/year for their study areas. Net retention of phosphorus in the studied riparian forest buffers was from 24%-­‐81% of input, or about 1 lb. of P/acre/year retained mainly from the retention of particulate phosphorous into the soils of the system. A study out of France estimated the nitrate buffering capabilities of a riparian forest over several seasons and noted a rate of up to 254 lbs. N/acre/year with a mean rate of 10.2 lbs. N/acre/year (Pinay 1993). The vegetation cover of the study sites included primary tree species of the Elm, Alder, and Ash families which are similar species to the composition of the riparian zones of forest in the Columbia Bottomlands (McFarlane -­‐-­‐-­‐-­‐). The study concluded that regardless of season, 30 meters or riparian 38 forest buffer strip along the course of the river or stream was enough to remove all the nitrates from entering the system through the groundwater (Pinay 1993). In a 1999 report for the Integrated Assessment on Hypoxia in the Gulf of Mexico, nitrogen loss rates are presented from several sources of literature and overall represent a removal rate range of 8 lbs. N/acre/year to 61 lbs. N/acre/year; all sources reported forested swamps and riparian zones as nitrogen sinks (Mitsch 1999). Mitsch reports that a study out of Georgia found a forested floodplain wetland removed 14% nitrogen through retention and 61 % through denitrification for a total of 75% removal, and the forest also had a 30% retention rate of phosphorus (Lowrance 1984). He also discusses a study in Iowa that found nitrate concentrations of 12 mg-­‐N/L in the groundwater of a field were reduced to 3 mg-­‐N/L through a forested riverine buffer (Mitsch 1999). Totals suspended solids removal of 40-­‐90%, total phosphorus removal of 30-­‐95%, and total nitrogen removal of 40-­‐70% are reported to result from water flowing through and over a riparian forest vegetated filter strip (Mitsch 1999). Riparian zone denitrification rates of 26-­‐35 lbs. N/acre/year were reported out of Carolina and sediment removal rates of up to 90% off of agricultural fields were also recorded (Naiman and Decamps 1997). Sheet flow runoff over agricultural fields and undeveloped, unchannelized natural areas facilitate sediment deposition in mature riparian forests that further contributes to the removal of sediment-­‐
bound pollutants from surface runoff (Naiman and Decamps 1997). The flat, low-­‐lying floodplain that maintains the topography of the Columbia Bottomlands is more conducive to sheet flow type runoff than the channelized riverine portions of the Colorado, San Bernard and Brazos Rivers. Overall, the shallow banks and lack of deep riverine systems as the water approaches the gulf encourages sedimentation of the floodplain which increases pollutant removal occurrences. The general range of rates reported from all of these studies along a riparian forest is consistent with, ĂŶĚƉŽƚĞŶƚŝĂůůLJŐƌĞĂƚĞƌƚŚĂŶ͕ƚŚĞƌĂŶŐĞŽĨƌĂƚĞƐƌĞƉŽƌƚĞĚĨƌŽŵĂƐƚƵĚLJĐŽŶĚƵĐƚĞĚŽŶĂ͞ĨŽƌĞƐƚĞĚ
ǁĞƚůĂŶĚ͟ŝŶƚŚĞDŝƐƐŝƐƐŝƉƉŝůůƵǀŝĂůsĂůůĞLJ, MAV, which reported low elevation forested wetland sites displayed denitrification rates of 25 lbs. Nitrogen/acre/ year, and high elevation forested sites displayed ĚĞŶŝƚƌŝĨŝĐĂƚŝŽŶƌĂƚĞƐŽĨϮ͘ϱůďƐ͘EͬĂĐƌĞͬLJĞĂƌ;:ĞŶŬŝŶƐĞƚĂůϮϬϬϵͿ͘:ĞŶŬŝŶƐ͛ƐƐƚƵĚLJŝŶƚŚe MAV also reported that through their findings, it was evident that as the wetland grows, the contribution of denitrification to the total removal volume of nitrate from the system increases from only 10% in the early years to nearly 50% after 90 years because of the change of growth rates of vegetation and soil sedimentation and deposition rates. During their study period, the total nitrate mitigated by the wetland bottomland forest, including retention and uptake, grew from 32 lbs. N/acre/year to 60 lbs. N/acre/year (Jenkins et al 2009). The similarities between the values found for denitrification and nitrogen and phosphorous retention of a forested riparian zone compared to a forested bottomland wetland indicate that the ĞĐŽƐLJƐƚĞŵĚĞƐŝŐŶĂƚŝŽŶĂƐĂ͞ƌŝƉĂƌŝĂŶĨŽƌĞƐƚ͕͟Ă͞ďŽƚƚŽŵůĂŶĚĨŽƌĞƐƚ͕͟ŽƌĂ͞ĨŽƌĞƐƚĞĚǁĞƚůĂŶĚ͕͟ǁŝůů
generally refer to very similar habitat types that exhibit similar functionalities for water quality improvement through denitrification and excess nutrient retention. 39 Table 8: Bottomland Forest Nutrient Removal/Retention Summary Table Source Lowrance et al (1997) NUTRIENT REMOVAL/RETENTION Location Rates of removal/retention of nutrients Georgia coastal plain riparian forest buffer system Pinay (1993) France Mitsch (1999) Iowa Gulf Coast vegetated strip Lowrance (1984) Georgia Naiman and Decamps (1997) Carolina Jenkins et al (2009) MAV Wetland Forest AVERAGES OF ALL STUDIES NITROGEN PHOSPHOROUS 27.6 lbs. N/acre/year 68-­‐74 lbs. N/acre/year total uptake 6.7-­‐8.9 lbs. P/acre/year total uptake 85-­‐90% nitrate input removal; 17-­‐35 lbs. N/acre/year 67-­‐89% total nitrogen retention; 23-­‐66 lbs. N/acre/year 24-­‐81% phosphorous input removal; 1 lbs. P/acre/year 10.2 lbs. N/acre/year 75% nitrate reduction through buffer 8-­‐61 lbs. N/acre/year 40-­‐90% total suspended solids removal 30-­‐95% phosphorous removal 40-­‐70% total nitrogen removal 75% nitrogen removal 30% phosphorous retention 26-­‐35 lbs. N/acre/year 90% removal of N off agricultural fields 25 lbs. N/acre/year low elevation forested sites 32-­‐60 lbs. N/acre/year 37.8 lbs. N/acre/year; 75.6% removal/retention 5.5 lbs./acre/year; 52% removal/retention Carbon Sequestration and Sink Capacity Tree species composition The tree species composition of the bottomlands is rather unique from coastal bottomland hardwood forests of other states, but with similar families and groups represented, still falls into a recognizable forest type as designated by the USDA and Society of American Foresters. The Columbia Bottomlands are dominated by roughly 87% basal area of canopy cover hardwood tree species (Rosen 2008). The dominant hardwood tree species seen in more frequently flooded areas are green ash, American elm, water hickory and less frequently flooded areas contain more sugarberry, cedar elm and water oak (Rosen 2008). A study conducted on three different tracts of land representing three different successional stages revealed that sugarberry, green ash, and American elm were the most dominant species present (McFarlane -­‐-­‐-­‐-­‐). Live oak and pecan are also found on the clay back flats of the area (Rosen 2008). The Society of American Foresters type 93 forest Sugarberry ʹ American elm ʹ Green Ash is the best represented forest type in the bottomlands, but little carbon data could be found for this specific of a forest classification (McFarlane -­‐-­‐-­‐-­‐). Dickson described the bottomland hardwoods of Texas as oak-­‐gum-­‐cypress forests in 1987 which is a well-­‐recognized national forest type (McFarlane -­‐-­‐-­‐-­‐). The Columbia Bottomlands also seem to fall in the forest type classification of elm-­‐ash-­‐cottonwood in the 40 more riverine areas and therefore will be considered a mix of both types in certain carbon sequestration data calculations for this report. Carbon accumulation and stock dŚĞĂŶŶƵĂů͞ĂĐĐƵŵƵůĂƚŝŽŶ͟ŽĨĐĂƌďŽŶƐƚŽƌĞĚǁŝƚŚŝŶĂĨŽƌĞƐƚĂŶĚƚŚĞƌĞƐƵůƚŝŶŐůŽŶŐ-­‐term storage carbon ͞ƐƚŽĐŬ͟ŽĨĨŽƌĞƐƚƐŚĂǀĞďĞĐŽŵĞǀĂůƵĂďůĞĐŽŵƉŽŶĞŶƚƐŝŶĂƐƐĞƐƐŝŶŐƚŚĞƚƌƵĞǀĂůƵĞĂŶĚĨƵŶĐƚŝŽŶĂůŝƚŝĞƐŽĨ
ecosystem services provided by them (Simpson 2013). The bottomland hardwood forests of the gulf coast not only store large masses of carbon in the high volumes of wetland grasses, understory vegetation, large hardwood tree and organic laden soils, but also sequester carbon at high rates each year through natural vegetation growth, soil formation, and biogeochemical processes within the wetland soils and waters. Clearing a forested area of above-­‐ground biomass releases stored carbon from the woody vegetation as it is burned or decomposed regardless if on-­‐site or off. Carbon dioxide and methane account for 80% of the global warming potential of all greenhouse gases meaning the release of these two gases would have a serious impact on global climate change rates. As bottomland wetlands are drained, the soils become aerobic and stored organic carbon is quickly oxidized to carbon dioxide; this changes the basic function of the land from a carbon sink as a wetland to a carbon source as dry land (Coastal Wetland Forest Conservation and Use Science Working Group 2005). The process can go both ways however; if historically-­‐forested bottomland habitat that has already been drained and cleared for agriculture or other uses were to be restored to its original ecosystem type, carbon dioxide would be removed from the atmosphere and sequestered into the forest biomass and soils at a rate well above the sequestration potential for crops and the hydrologic regime of the area can begin to reform (Jenkins 2009). It has been noted that wetlands do produce other greenhouse gases such as methane and nitrous oxide, but their influence is easily countered and surpassed by the rate of carbon sequestration of a growing wetland forest such as the Columbia Bottomlands (Jenkins 2009). Jenkins reports in a study about wetland restoration in the MAV that forested Wetland Restoration Project sites of their study sequester up to 82.2 Mg of new CO2 per hectare per 5 year period which is approximately equal to 2 tons C/acre/year5. Based off data displayed in the report by Jenkins et al, total carbon storage of an afforested bottomland hardwood forest, noted as an oak-­‐gum-­‐cypress forest type, would range from about 11 tons C/acre (25 Mg C/ha) in the first years increasing to about 102 tons C/acre (230 Mg C/ha) after 90 years6. These values represent total carbon accumulation of live biomass, soil carbon, and other carbon (litter and dead trees). The age of the Columbia Bottomlands, based off historical accounts of the area, is well over a century and therefore would currently contain an estimated 115 tons C/acre based on this analysis (Jenkins 2009). A similar study out of the Lower Mississippi Valley also estimated the carbon storage of bottomland hardwood forests and reported similar results to Jenkins ʹ reporting only the total live tree biomass at about 10 tons C/acre at 10 years and smoothing out at around 90 tons C/acre after 90 years (Shoch 2009). If you consider only the total live tree biomass accumulation in the report from Jenkins, ShŽĐŚ͛ƐǀĂůƵĞƐĂƌĞconsistent. These values lead to an estimated total of 13.5-­‐17.25 Million tons of carbon in the Columbia Bottomlands7. 5
One ton of carbon is equal to a weight of 3.67 tons when in the form of carbon dioxide (3 ton C = 11 ton CO2). For the purposes of this report, all carbon values are converted or reported as imperial tons of elemental carbon, C. 1 hectare =2.47 acres, 1 Mg = 1.102 imperial tons 6
Values were obtained through examination of a graph within the report and not assumed to be 100% accurate, a direct conversion was made from the estimated value of Mg C/ha to tons C/acre to obtain the values discussed in this report. 7
The estimated biomass of 90-­‐115 tons C/acre multiplied by the estimated 150,000 acres of remaining bottomland forest in the Columbia Bottomlands 41 A 2002 study by NFWF and Winrock International predicts that the complete conservation of the remaining area of Columbia Bottomlands will prevent the release of 86 tons of C/acre into the ĂƚŵŽƐƉŚĞƌĞ;ĞůĂŶĞLJĞƚĂůϮϬϬϮͿ͘ƐƚƵĚLJŽĨ>ŽƵŝƐŝĂŶĂ͛ƐĐŽĂƐƚĂůǁĞƚůĂŶĚĨŽƌĞƐƚƐƌĞƉŽƌƚƐƚŚĞƐŽŝůĐĂƌďŽŶ
density of a wetland forest to be 201 tons/acre and the soil carbon density of an upland forest to be 40 tons/acre (Trettin and Jurgensen 2003). If the bottomlands are conservatively estimated from this study to hold 40 tons C/acre in the soil, added with the NFWF estimation of 86 tons C/acre, the 150,000 acres of remaining bottomlands are estimated to contain 126 tons C/acre of bottomland habitat (total biomass) which would yield a total estimate of 18.9 million tons C held within the Columbia Bottomlands and would be released if the area is deforested (U.S. Fish and Wildlife Service 2013). Carbon densities by forest type are also reported by Heath et al 2003: for oak-­‐gum-­‐cypress forest types, a total forest carbon density of 115.8 tons C/acre and for elm-­‐ash-­‐cottonwood, 97 tons C/acre. These values lead to a calculation of 50/50 forest types over the estimated remaining acreage of the Columbia Bottomlands to be 15.96 million tons of total carbon. These values are in line with other estimations based on as 50/50 mixture of both FIA forest types. A US Department of Agriculture report by the Forest Service on carbon storage and accumulation rates in United States forest ecosystems reports the average carbon accumulation in trees of oak-­‐gum-­‐cypress forests and elm-­‐ash-­‐cottonwood forests of the South Central region to be 1685 lbs./acre/yr. to 1999 lbs./acre/year respectively based on values reported from 1987 (Birdsey 1992). Assuming the Columbia Bottomlands are 50% of each forest type, the rate of carbon accumulation in the biomass of the forest is 1842 lbs. C/acre/year. More recent estimates from the Texas Statewide Assessment of Forest Ecosystem Service conducted by the Texas A&M Forest Service in 2013 classifies a Hardwood-­‐Bottomland forest type as an equivalent FIA oak-­‐gum-­‐cypress type group and the Hardwood-­‐Riparian forest type as an equivalent elm-­‐as-­‐cottonwood FIA forest type group. Texas A&M Forest Service reports the above ground live-­‐carbon accumulation rates for the hardwood-­‐bottomland and riparian forest types as 1.12 tons C/acre/year and 1.21 tons C/acre/year (Simpson 2013)8. Using the same 50/50 forest-­‐type assumption as before, the Columbia Bottomlands would accumulate carbon at a rate or 2330 lbs./acre/year. These calculations yield a mean value of 2086 lbs./acre/year or 1.04 tons C/acre/year ĂĐĐƵŵƵůĂƚŝŽŶŝŶƚŚĞůŝǀĞďŝŽŵĂƐƐŽĨƚŚĞŽůƵŵďŝĂŽƚƚŽŵůĂŶĚ͛ƐŚĂƌĚǁŽŽĚĂŶĚƌŝƉĂƌŝĂŶĨŽƌĞƐƚƐ͘ The Texas Statewide Assessment of Forest Ecosystem Services also analyses carbon stock and accumulation by categorizing forests by ecoregion and forest type together and reports that hardwood-­‐
ďŽƚƚŽŵůĂŶĚĨŽƌĞƐƚƐŽĨƚŚĞ͞ĐŽĂƐƚĂůǁŽŽĚůĂŶĚƐ͟ƌĞŐŝŽŶĂĐĐƵŵƵůĂƚĞĐĂƌďŽŶĂƚĂƌĂƚĞŽĨϮϵϭ͕ϭϬϬƚŽŶƐ
C/year and the hardwood-­‐riparian forests of the coastal woodlands accumulate 427,800 tons C/ year yielding a 50/50 average of 359,450 tons C/year for the hardwood forests of the Columbia Bottomlands (Simpson 2013). Dividing that rate by 150,000 acres yields a rate of 2.4 tons C/acre/year. This value is significantly higher (roughly double) than what was estimated from the USDA and FIA rates discussed ƉƌĞǀŝŽƵƐůLJĂŶĚŝƐůŝŬĞůLJďĞĐĂƵƐĞƚŚĞ͞ĐŽĂƐƚĂůǁŽŽĚůĂŶĚƐ͟ƌĞŐŝŽŶŽĨƚŚŝƐƐƚƵĚLJĂůƐŽĞŶĐŽŵƉĂƐƐĞƐŽƚŚĞƌ
bottomland habitat further south that is not considered as part of the Columbia Bottomlands. 8
The report states that these are conservative estimates of accumulation rates and that actual rates are likely much higher 42 Table 9: Bottomland Forest Carbon Services Summary Table Source Jenkins er al (2009) Scoch (2009) Delaney (2002) CARBON SEQUESTRATION AND STOCK Location Sequestration rates and stock values Mississippi Alluvial Valley Mississippi Alluvial Valley Columbia Bottomlands Trettin and Jurgansen (2003) Heath et al (2003) Louisiana coastal forest USDA (1992) Simpson (2013) AVERAGES OF ALL STUDIES forest type estimates forest type estimates ecoregion estimate ACCUMULATION forest type estimates STOCK TOTAL BIOMASS 2 tons C/acre/year accumulation 115 tons C/acre total biomass stock 90 tons C/acre live tree biomass stock 86 tons C/acre live biomass stock 40-­‐201 tons C/acre soil carbon density 106.4 tons C/acre total biomass stock Columbia Bottomlands Carbon Stock estimate 18.9 million tons C total biomass stock 15.96 million tons C total biomass stock 17.25 million tons C total biomass stock 359,450 tons C/year total biomass accumulation 1842 lbs. C/acre/year accumulation 2330 lbs. C/acre/year accumulation 2.4 tons C/acre/year accumulation 1.8 tons C/acre/year 109 tons C/acre 17.37 million tons C in Columbia Bottomlands Migratory Bird Habitat and the Birding Industry The Columbia Bottomlands represent the last extent of coastal bottomland hardwood and riverine forested wetlands along the Texas coast of the Gulf of Mexico. Each spring, neotropical songbirds migrate from their wintering grounds in Central and South America to North America (Figure 16). After last setting foot on land in the Yucatán Peninsula, the bottomland forests of the Central Texas Coast offer a vital refuge to these migratory birds as they complete the long and exhausting journey over the Gulf of Mexico (Faulkner 2004). The Columbia Bottomlands offer food, water, shelter, and resting places for millions of individuals as they replenish their energy stores after the tiring flight over the gulf. A study conducted by the US Fish and Wildlife Service on the Columbia Bottomlands reported an observed 237 species of birds compromising over 29 million individuals that utilize the bottomland forests during their annual migration (US Fish and Wildlife Service 1997). The massive populations of birds can be seen on radar, flocking into the Columbia bottomlands and taking flight further north after a short recovery and resting period (Gauthreaux 2002). Because this ecosystem is the only remaining expanse of forest adjacent to the Gulf of Mexico, the Columbia Bottomlands are a critical stopover spot for these Neotropical migrants and have thus attracted thousands of wildlife viewers, avitourists, and birding enthusiasts to the area every year to observe this unique and dense collection of species. The forests are also home to millions of resident birds that spend all 12 months of the year within the 43 bottomland forests and along the various rivers and streams. Waterfowl also compromise a large number of the resident and migratory bird species and areas of the bottomlands that are open to duck hunting, primarily near the coast, provide a valuable economic resource that is accessible to the avid hunting population of Texas. Bird watching, or birding, has recently gained momentum as a popular form of outdoor recreation and ecotourism. The number of Unites States residents that participate in bird watching grew from 21 million in 1983 to 69 million in 2000 with a collective 33.5% of the adult population of the US that reports viewing and photographing birds (National Survey on Recreation and the Environment 2000). In-­‐state expenditures of US wildlife watchers increased by 12% from 2001 to 2011 (USDOI 2011). Birders spent an estimated $15 billion on their trips and $26 billion on equipment in 2011 across the US, an average of $825 per person over the course of the year on birding equipment and trips designated for the purpose of bird watching (US Fish and Wildlife Service 2013). Birding Source: http://lyndagoff.com/songbird-­‐migration-­‐across-­‐the-­‐gulf-­‐of-­‐mexico/ expenditures in 2011 created 666,000 jobs and $31 million in employment F IGURE 16: NEOTROPICAL MIGRATION ROUTES INTERSECTING income $6 billion in State tax revenue and OVER THE C OLUMBIA B OTTOMLANDS $7 billion in Federal tax revenue (US Fish and Wildlife Service 2013). The birding industry in the US is obviously a huge economic asset to the states and communities that support, encourage, and provide avitourism opportunities. A USFWS report on the birding rates and economic analyses of birding in the United States reports that Texas is home to 2,238,000 birders with 95% of them state residents (US Fish and Wildlife Service 2013). dŚĞŵĞĂŶŶƵŵďĞƌŽĨĚĂLJƐƐƉĞŶƚďŝƌĚǁĂƚĐŚŝŶŐďLJĂ͞ďŝƌĚĞƌ͟ŝŶdĞdžĂƐǁĂƐŶŽƚĞĚĂƐϭϯϮĚĂLJƐŽƵƚŽĨƚŚĞ
year, above the US average of 110 days (US Fish and Wildlife Service 2013). The Great Texas Coastal Birding Trail brings avitourists from all over the state and country, some even internationally to the central Texas coast to view the large variety of bird species. Of people surveyed who traveled the central coast portion of the Great Texas Birding Trail, the average traveler spend 31 days of the year viewing wildlife on the trail and spent an average of $78 per day while on the trail (Eubanks 1999). Visitors to the High Island Audubon Sanctuary, part of the Upper Texas Coast portion of the trail spent contributed at least $2.5 million to the economy of the area in 1992, a report estimates (Eubanks 1993). Nature tourism development in rural communities, such as those surrounding the Columbia Bottomlands, diversifies local economies and provides benefits to the people who live there such as economic boosts (local tax revenues were $301 million in 2001) and instilling pride in community and providing jobs for residents (Travel Facts 2003). The nature-­‐tourism industry is a powerful economic driver that can capitalize easily on the promotion of prime bird habitat provided by the bottomland forests of the central Texas coast. Millions of dollars are spent annually on birding trips, supplies, and education materials. There is no doubt that the bird habitat provided by the Columbia Bottomlands is an ecosystem service that provides economic revenue and other tourism-­‐based benefits and economic 44 support to the local communities and environmental-­‐agencies that promote the conservation and sustainable use of the bottomland forests of the central coast. T ABLE 10: BOTTOMLAND F OREST B IRD HABITAT VALUES Source BIRDING INDUSTRY Study Reference Values USFWS (1997) Columbia Bottomlands 237 species; 29 Million birds National Survey on Nationally 69 million bird-­‐watchers Recreation and the 33.5% view and photograph birds Environment (2000) USFWS (2013) Nationally $825 per person on birding trips industry created 666,000 jobs $31 million employment income $6 billion State taxes $7 billion Federal taxes Texas 2,238,000 birders 132 days/year spent birding Eubanks (1993) Coastal Birding Trail 31 days/year traveling on trail $78 spent per day on trail High Island $2.5 million contribution to economy Travel Facts (2003) local communities $301 million tax revenues in 2001 There are many issues surrounding the preservation of the bottomlands as a pristine neotropical migrant bird habitat, therefore threatening the economic services provided by bird-­‐watching, outdoor recreation, and eco-­‐tourism revenues in the area. The treat of fragmentation by urban expansion and residential development has already been discussed, along with agricultural use of land that is ĚĞĨŽƌĞƐƚĞĚĂŶĚĂůƚĞƌĂƚŝŽŶŽĨƚŚĞŶĂƚƵƌĂůŚLJĚƌŽůŽŐŝĐƌĞŐŝŵĞƚŚĂƚƐƵƐƚĂŝŶƐƚŚĞďŽƚƚŽŵůĂŶĚ͛ƐƵŶŝƋƵĞ
species composition and forest structure. The diversity of tree and flowering plant species supports the wide diversity of bird species and other animal species that use the bottomlands as temporary or permanent habitat. A disturbed forest habitat will lack the natural structure and diversity that is maintained by the long-­‐standing natural forest stands, and therefore will provide a less-­‐suitable habitat for many species. Changes in understory habitat can affect the foraging success of birds that have specialized diets for certain berries, flower nectar, and flowering plant fruits (Faulkner 2004). Even a slight change in forest composition, understory diversity, or hydrologic patterns can affect the stands suitability for use of many different species, ultimately affecting the long-­‐term viability of the habitat which leads to loss of bird diversity. Even over grazing by native species such as white-­‐tailed deer can affect the understory composition of the forest and make it less-­‐suitable for migratory bird species that depend on food groups found in the understory layers of the bottomland forests (Barrow 2000). Fragmentation of the forest not only removes forested land but also increases the amount of forest edge, making the intact portions more vulnerable to invasive species infiltration. Chinese Tallow, Sapium sebiferum, is an invasive tree species that is fast growing, hardy, and tolerant of a wide range of environmental and soil conditions and has thus overrun native vegetation across many ecosystems in 45 the coastal states, including coastal prairies http://www.refugefriends.org/photos/var/resizes/Gallery-­‐-­‐-­‐
Aves/Passeriformes-­‐-­‐-­‐
and bottomland forests. While Tallow can still Other/Baltimore_Oriole_005.jpg?m=1369268888
be used as cover habitat by most migratory and resident bird species, the invasive species does not support the native food sources that a lot of migrant birds depend on such as native species of caterpillars (Barrow and Renne 2001). This seemingly insignificant change in food source availability caused by the domination of Chinese Tallow plants over native tree species that do support this food source can cause a dramatic shift in the food web and affect the availability of other food sources used by different species. F IGURE 17: BEAUTIFUL BIRDS SUCH AS THIS B ALTIMORE The migrant and resident bird habitat that is O RIOLE CAN BE SEEN EACH YEAR AS THEY MIGRATE provided by the coastal bottomland hardwood THROUGH THE C OLUMBIA B OTTOMLANDS forests of Texas is a significant ecosystem service that provides huge economic and cultural benefits to surrounding communities and the entire gulf-­‐coast. Without such ecologically valuable and rare habitat available to them along their journey, neo-­‐tropical migrant populations would dwindle, limiting the economic and recreation benefits that we all procure from the birding and wildlife-­‐
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opportunities for our residents and avitourists, but the vital habitat provided by an intact coastal bottomland forest secures the proliferation of bird populations that benefit every other state that these migrants migrate through. References Barrow, W. C., Jr., and I. Renne. 2001. Interactions between migrant land birds and an invasive exotic plant: The Chinese tallow tree. Flyway 8:11. Barrow, W. C., Jr., C. Chen, R. B. Hamilton, K. Ouchley, and T. J. Spengler. 2000. Disruption and restoration of en route habitat, a case study: The Chenier Plain. In: F. R. Moore, editor. Stopover ecology of nearctic-­‐neotropical land bird migrants: Habitat relations and conservation implications. Studies in Avian Biology 20: 71-­‐87. Barrow, W. C., L. A. Johnson-­‐Randall, M. S. Woodrey, J. Cox, E. Ruelas I., C. M. Riley, R. B. Hamilton, and C. Eberly. 2005. Coastal forests of the Gulf of Mexico: a description and thoughts on their conservation. Bird Conservation Implementation and Integration in the Americas: Proceedings of the Third International Partners in Flight Conference, C. J. Ralph and T. D. Rich, editors pp. 450-­‐464. Berg, R. 2009. Tropical Cyclone Report Hurricane Ike (AL092008) 1-­‐14 September 2008. National Hurricane Center, 1-­‐41. 46 Birdsey, R. A. (1992). Carbon storage and accumulation in United States forest ecosystems. US
Department of Agriculture, Forest Service, General Technical Report WO-59 Coastal Wetland Forest Conservation and Use Science Working Group. 2005. Conservation, Protection ĂŶĚhƚŝůŝnjĂƚŝŽŶŽĨ>ŽƵŝƐŝĂŶĂ͛ƐŽĂƐƚĂůtĞƚůĂŶĚ&ŽƌĞƐƚƐ͗&ŝŶĂůZĞƉŽƌƚƚŽƚŚĞ'ŽǀĞƌŶŽƌŽĨ>ŽƵŝƐŝĂŶĂ͘
Jim L. Chambers, School of Renewable Natural Resources, Louisiana State University Agricultural Center. ŽƐƚĂŶnjĂ͕Z͕͘Ě͛ƌŐĞ͕Z͕͘ĚĞ'ƌŽŽƚ͕Z͕͘&ĂƌďĞƌ͕^͕͘'ƌĂƐƐŽ͕D͕͘,ĂŶŶŽŶ͕͕͙͘ǀĂŶĚĞŶĞůƚ͕D͘;ϭϵϵϳͿ͘dŚĞ
ǀĂůƵĞŽĨƚŚĞǁŽƌůĚ͛ƐĞĐŽƐLJƐƚĞŵƐĞƌǀŝĐĞƐĂŶĚŶĂƚural capital. Nature, Vol 387, 253-­‐260. Dawson, Bill. 2004, July 1. Columbia Bottomlands: It's not just what rivers are, it's what they create. Texas Parks and Wildlife Magazine. Day Jr, John W. and Hunter, Rachael G. 2013. Restoration and Conservation Of Coastal Forested Wetlands in The Gulf Of Mexico: A Report Prepared For U.S. Endowment for Forestry and Communities. Greenville, SC. Delaney M., S. Brown and D. Shoch. 2002. The Carbon content of Austin Woods/Columbia bottomland hardwood forests, Winrock International, Arlington VA. Eubanks, T., Kerlinger, P. and R.H. Payne. 1993. High Island, Texas: Case Study in Avitourism. Birding, Vol. XXV: Number 6. American Birding Association, Inc. Eubanks, Ted and John Stoll. 1999. Avitourism in Texas: Two Studies of Birders in Texas and Their Potential Support for the Proposed World Birding Center. Texas Parks and Wildlife Contract No. 44467. Faulkner, S. 2004. Urbanization Impacts on the Structure and Function of Forested Wetlands. Urban Ecosystems, 7, 89-­‐106. Faulkner, S.P., Barrow, W., Keeland, B., Walls, S., Moorman, T., Twedt, D., et al., 2008. Assessment of Ecological Services Derived from U.S. Department of Agriculture Conservation Programs in the Mississippi Alluvial Valley: Regional Estimates and Functional Condition Indicator Models. Interim Report. USDA NRCS Gauthreaux, Jr. Sidney A. 2002. Radar Ornithology and Bird Conservation. Available at: http://www.gcbo.org/. Accessed March 27, 2012. Hendrickson, O. Q., Jr. 1981 Flux of nitrogen and carbon gasses in bottomland soils of an agricultural watershed. PhD dissertation. University of Georgia, Athens, 210 pp. (Dissertation Abstract 82-­‐
1544). Hunter, R.G. and Faulkner, S.P. 2001. Denitrification potentials in restored and natural bottomland hardwood wetlands. Soil Sci. Soc. Am. J. 65, 1865ʹ1872. 47 Hunter, R.G. and S.P. Faulkner. 2001. Denitrification potentials in restored and natural bottomland hardwood wetlands. Soil Science Society American Journal 65:1865-­‐1872. Jenkins, W., Murray, B., Kramer, R., & Faulkner, S. 2009. Valuing ecosystem services from wetlands restoration in the Mississippi Alluvial Valley. Ecological Economics, 69, pp 1051-­‐1061. Lowrance, R. R., R. L. Todd, J. Fail, Jr., O. Hendrickson, Jr., R. Leonard, and L. Asmussen. 1984. Riparian forests as nutrient filters in agricultural watersheds. BioScience 34:374ʹ377 Lowrance R, Altier LS, Newbold JD, Schnabel RR, Groffman PM, et al. 1997. Water quality functions of Riparian forest buffers in Chesapeake Bay watersheds. Environmental Management 21: 687ʹ
712. doi: 10.1007/s002679900060 Masters, Jeffrey. 2014. U.S. Storm Surge Records. Weather Underground. http://www.wunderground.com/hurricane/surge_us_records.asp? McFarlane, Dr. Robert W. -­‐-­‐-­‐-­‐. A Comparison of Three Brazos River Forests. Millennium Ecosystem Assessment. 2005. Ecosystem and human well-­‐being: current state and trends. Washington DC: Island Press. Miller, Wesley L. and Amanda L. Bragg. 2007. Soil Characterization and Hydrological Monitoring Project, Brazoria County, Texas, Bottomland Hardwood Vertisols. United States Department of Agriculture, Natural Resources Conservation Service, Temple, Texas Mitsch, W.J., J.W. Day, Jr., J.W. Gilliam, P.M. Groffman, D.L. Hey, G.W Randall, and N. Wang. 2001. Reducing nitrogen loading to the Gulf of Mexico from the Mississippi River Basin: Strategies to counter a persistent ecological problem. BioScience 51:373-­‐388. Mitsch, William J., John W. Day, Jr., J. Wendell Gilliam, Peter M. Groffman, Donald L. Hey, Gyles W. Randall, and Naiming Wang. 1999. Reducing Nutrient Loads, Especially NitrateʹNitrogen, to Surface Water, Ground Water, and the Gulf of Mexico: Topic 5 Report for the Integrated Assessment on Hypoxia in the Gulf of Mexico. NOAA Coastal Ocean Program Decision Analysis Series No. 19. NOAA Coastal Ocean Program, Silver Spring, MD. 111 pp. Naiman, Robert J. and Decamps, Henri. 1997. The Ecology of Interfaces: Riparian Zones. Annual Review of Ecology and Systematics, Vol. 28, (1997), pp. 621-­‐658 Accessed at: http://www.jstor.org/stable/2952507 National Survey on Recreation and the Environment 2000. USDA Forest Service and USDC, NOAA Pinay, G., Roques, L., & Fabre, A. 1993. Spatial and Temporal Patterns of Denitrification in a Riparian Forest. The Journal of Applied Ecology, 30(4), 581-­‐581. Retrieved September 24, 2014, from http://www.jstor.org/stable/2404238 Rosen, D. J., De Steven, D., & Lange, M. L. 2008. Conservation strategies and vegetation
characterization in the Columbia Bottomlands, an under-recognized southern floodplain forest
formation. Natural Areas Journal, 28(1), 74-82.
48 Shoch, D. T., Kaster, G., Hohl, A., & Souter, R. (2009). Carbon storage of bottomland hardwood
afforestation in the Lower Mississippi Valley, USA. Wetlands, 29(2), 535-542. Simpson, H., Taylor, E., Li, Y., & Barber, B. (2013). TEXAS STATEWIDE ASSESSMENT OF FOREST ECOSYSTEM SERVICES: A comprehensive analysis of regulating and cultural services provided by dĞdžĂƐ͛ĨŽƌĞƐƚƐ͘dĞdžĂƐΘD&ŽƌĞƐƚ^ĞƌǀŝĐĞ͘ŽůůĞŐĞ^ƚĂƚŝŽŶ͕dy͘ Sun, G., Mcnulty, S., Amatya, D., Skaggs, R., Swift, L., Shepard, J., & Riekerk, H. 2002. A comparison of the watershed hydrology of coastal forested wetlands and the mountainous uplands in the Southern US. Journal of Hydrology, (263), 92-­‐104. Travel Facts: A Quick Reference Guide to Current Travel Facts and Trends, Texas Economic Development, 2003 Trettin, C.C., and M.F. Jurgensen. 2003. Carbon cycling in wetland forest soils. Pages 311-­‐331 in J.M Kimble, et al. (eds.). The Potential of U.S. Forest Soils to Sequester Carbon and Mitigate the Greenhouse Effect. CRC Press, Boca Raton, FL. U.S. Department of the Interior, U.S. Fish and Wildlife Service, and U.S. Department of Commerce, U.S. Census Bureau. 2011 National Survey of Fishing, Hunting, and Wildlife-­‐Associated Recreation. h͘^͘&ŝƐŚĂŶĚtŝůĚůŝĨĞ^ĞƌǀŝĐĞ͘ϭϵϵϳ͘&ŝŶĂůƉƌŽƉŽƐĞĚƵƐƚŝŶ͛ƐǁŽŽĚƐĐŽŶƐĞƌǀĂƚŝŽŶƉůĂŶ͕ůĂŶĚƉƌŽƚĞĐƚŝŽŶ
comƉůŝĂŶĐĞĚŽĐƵŵĞŶƚĂŶĚĐŽŶĐĞƉƚƵĂůŵĂŶĂŐĞŵĞŶƚƉůĂŶ͗ƵƐƚŝŶ͛ƐǁŽŽĚƐƵŶŝƚƐŽĨƚŚĞƌĂnjŽƌŝĂ
National Wildlife Refuge complex. Albuquerque, NM: Fish and Wildlife Service, U. S. Department of the Interior. h͘^͘&ŝƐŚĂŶĚtŝůĚůŝĨĞ^ĞƌǀŝĐĞ͘ϮϬϭϯ͘dĞdžĂƐDŝĚͲĐŽĂƐƚEtZComplex Comprehensive Conservation Plan and Environmental Assessment: Appendix I: Land Protection Plan. Brazoria, TX: Texas Mid-­‐coast National Wildlife Refuge Complex. Albuquerque, NM: US Fish and Wildlife Service, U.S. Department of the Interior. US Fish and Wildlife Service. 2013. Birding in the United States: A Demographic and Economic Analysis. U.S. Fish and Wildlife Service, Arlington, VA. 49 Texas Coastal Prairies Prairies and grasslands along the Texas Gulf Coast provide a wealth of resources and services to Texas residents in the form of climate regulation, water quality, bird habitat, and nutrient cycling. These services are often under-­‐appreciated since they are public goods, and not readily quantified. Texas coastal prairies have been lost at an alarming rate due to land conversion and development. Once covering over 6.5 million acres of Texas land, prairies now occupy less than 1% of these lands ʹ or only 65,000 acres (Baldwin et al, 2007). Efforts have been made over the last several decades to estimate the value of prairie ecosystem services in order to more accurately assess the importance of Texas coastal prairies, and stem the loss of prairie land. In this paper, relevant literature is summarized regarding carbon sequestration, bird habitat, water regulation, and nutrient cycling, in order to give decision-­‐
makers a clearer view of the vital role of prairie ecosystems. Carbon Sequestration Prairies have the potential to store large amounts of carbon depending on land management practices and vegetation cover. Native prairie grasses have extensive root systems that can go as deep as 15 feet underground (Figure 18) ; carbon is stored both in the root systems of these plants, as well as the soil underground as plants grow and form new soil (Jones & Donnelly, 2004). Studies have shown that natural prairie and grassland ecosystems hold much more carbon in their soils than agricultural lands, and one study estimates that 5000 million metric tons of carbon has been released into the air from the conversion of natural land to agricultural land in the U.S. (Lal et al 1999). For this reason, U.S. government programs, like the Conservation Reserve Program, have encouraged the transition of agricultural lands back to natural grasslands. Several studies have shown that this practice can lead to a vast increase in soil organic carbon levels at these sites. The reason prairies and grasslands are able to sequester more carbon than agricultural land are numerous, but the main factors are that prairie and grassland plants provide a greater supply of carbon to the soil underground and that natural grasslands have an http://il-­‐elgin3.civicplus.com/images/pages/N1279/image007.gif increased carbon F IGURE 18: R OOT SYSTEMS OF COMMON PRAIRIE PLANTS 50 residence time because the need for tilling or other major soil disturbances are not present (Jones & Donnelly, 2004). A study by Potter et al (1999) aimed to calculate the soil organic carbon levels and the rate of carbon sequestration at sites where degraded agricultural land had been restored to natural grassland. The study compared sites that had been returned to grassland 6, 26, and 60 years ago, a site that had been used for agriculture continuously for 100 years, and a site of pristine prairie that had never been used for agriculture. Three different study areas were chosen, each with a pristine prairie, a restored prairie, and agricultural land. Two of the areas were located near Temple, TX, and the third was located near Riesel, TX. Results from the analysis showed that soil organic carbon was highest in the native prairies soils, lowest in the agricultural land, and somewhere in the middle for the restored prairie. Using a relationship between the time since the prairie was restored, and the amount of organic carbon in the soil, researchers developed a linear function to model the carbon sequestration rates. It was found that restored grasslands could sequester 428 lbs. C per acre per year. Another study conducted by Sims & Bradford (2001) compared the carbon sequestration ability of native prairie grass to sagebrush. Monitoring sites were chosen in the Southern Plains prairie, and CO2 fluxes from each vegetation site were monitored over two years. The results of this paper reported that the grass site was able to sequester an average of 623 lb. C per acre per year, and the sagebrush site sequestered an average of only 58 lb. C per acre per year. Since water use and availability was similar for each vegetation type, researchers concluded that prairie grass is a more viable carbon sink than sagebrush. Though not conclusive, this study is enlightening to the prevalent scientific debate about the effects of woody plant encroachment in grasslands on carbon pools. The carbon dioxide fluxes of an ecosystem depend on several climatic factors such as season and moisture. Suyker & Verma (2001) studied the net ecosystem CO2 exchange (NEE) for a prairie in north-­‐
central Oklahoma over a 20 month period. The aim of this paper was to quantify annual carbon exchanges, and understand how environmental factors such as land use, management, climate and nutrient levels influence carbon flows. Results showed that carbon sequestration in prairies is heavily dependent on season. During the winter months, carbon flux was negligible, but increased substantially in spring and summer. By the end of April, NEE was 0.25 mg CO2 per m2 per second, and by July the NEE was 1.4 mg CO2 per m2 per second. This number is consistent with other studies that have calculated similar NEE peaks: Ham and Knapp (1998) calculated 1.0 mg CO2 per m2 per second for August, Kim and Verma (1990) calculated a seasonal peak of 1.3 mg CO2 per m2 per second, and Dugas et al (1999) calculated 1.2 mg CO2 per m2 per second for the summertime peak. This study also demonstrated that soil moisture also influences carbon fluxes. During the dry season the NEE magnitude was only 0.2 mg CO2/m2/second, while it averaged 1.1 mg CO2/m2/second during the rainy period in August. It is noted in this study that during days that were very hot and dry, the flux changed signs and the prairie actually became a source for carbon as plants were being decomposed at higher rates than growth (Suyker and Verma 2001). After comparing the daytime and nighttime fluxes across all seasons, it was found that daytime carbon sequestration was 7154 lb. C/acre/year, and night 51 releases were 4768 lb. C/acre/year, yielding a net sequestration rate of 2386 lb. C/acre/year (Suyker and Verma 2001). Table 11: Prairie Carbon Services Summary Table Carbon Sequestration Rates Summary Table Study Suyker, Verma Year Location 2001 Oklahoma Prairie Type tallgrass Reported Seq. Rate seq. rate (lbs. C/ac/y) (g C/m2/y) 268 2386 Notes Monitoring conducted in 1996-­‐1998 Dugas et al 1998 Texas tallgrass 80 712 Sim, Bradford 2001 Southern Plains mixed-­‐grass 70 623 Bowen ratio/energy balance used Potter et al 1999 Texas restored grassland 48 428 Based on comparisons to a pristine prairie and degraded agricultural soil 117 1037 AVERAGES FROM ALL STUDIES: Nutrient Cycling It is important to understand how prairies provide nutrient cycling as a regulating service along the Texas Gulf Coast. Prairie vegetation, wetlands, and streams can uptake nutrients and modify them, which affects the flow rates and magnitudes of nutrients reaching downstream ecosystems (Tate, 1990). The presence of these nutrients in turn affects ecological processes occurring in downstream ecosystems; the coastal ecosystems of Texas are all connected in this way. Coastal prairies are unique because they are characterized by their small depressions and catchments, which make up 40% of the land around Galveston Bay (Enwright et al, 2011). Coastal prairie wetlands (CPWs) are significant sinks for inorganic nitrogen and phosphorus, and by capturing and controlling the release of nutrients, they help regulate water quality in Galveston Bay. However, in recent decades CPWs have been lost or degraded at a concerning rate, and presently over 98% of these wetlands along the Texas coast have been converted to agricultural lands or uplands (Forbes & Doyle, 2012). 52 Forbes et al (2012) conducted a study that monitored six coastal prairie wetlands and calculated the nutrient retention of each one. The study generated a mass balance using nutrient concentrations of incoming water, and comparing that to surface water nutrient levels in the wetland. Results found that there was significant variability in nutrient retention between sites and season, though there was no observable seasonal trend. However, among all six sites the prairie wetlands retained a significant amount of total incoming nitrogen. On average, the wetlands retained 826 kg N per km2 per year (7.36 lbs./acre/year), which represents 80.1% of total incoming nitrogen. Wetlands also filtered incoming phosphorus at a rate of 60 kg per km2 per year (0.54 lbs./acre/year), which corresponded to a 75% retention rate. The water quality regulation provided by CPWs is severely under-­‐valued by decision-­‐makers in Texas. dŚĞƐĞǁĞƚůĂŶĚƐĂƌĞĐŽŶƐŝĚĞƌĞĚƚŽďĞ͞ŐĞŽŐƌĂƉŚŝĐĂůůLJŝƐŽůĂƚĞĚ͟ĨƌŽŵƚŚĞ'ĂůǀĞƐƚŽŶĂLJƐLJƐƚĞŵ͕ĂŶĚĂƐ
such are not federally protected. Instead, each wetland is viewed as an individual unit, a method that does not take into account the cumulative regulating effect of the entire wetland prairie system. Without these depressional prairie wetlands, significantly elevated nutrient levels would reach Galveston Bay and affect commercial and recreational activities that depend on a healthy bay ecosystem. (Enwright et al, 2011). Prairie tallgrass also has the potential to store and cycle nutrients. Nitrogen content of a tallgrass prairie in northeastern Oklahoma was monitored over a several year period by Risser et al (1982). Data was collected for living biomass, dead biomass, and soils on the prairie. Using the nitrogen concentrations in each of these components, researchers calculate that native prairies contain between 200-­‐1550 g N per m2; there is variability in these numbers due to seasonal variations in N levels in the plant biomass. In addition, it was calculated that prairie grasses could remove 2.5 g N per m2 per year (22 lbs./acre/year), and transfer this nitrogen to the soil. Another study conducted in Missouri in 1969 found similar results in terms of nitrogen retention in prairie soils. This study calculated that prairie grass filtered 3.7 g per m2 per year (33 lbs./acre/year) of nitrogen though the shoot and root system of the prairie grass (Risser 1982). Seastedt (1988) attempted to determine the viability of different tallgrass components as nitrogen and phosphorus sinks. The study monitored a tallgrass prairie in Kansas, and measured the nitrogen and phosphorus levels in the stems and roots of the prairie grass. In addition, decomposition and mineralization rates of the plant biomass were measured. Nitrogen uptake rates were 1.94 g per m2 per year (17.3 lbs./acre/year), and phosphorus rates were 0.19 g per m2 per year (1.7 lbs./acre/year). Based on nutrient levels in the shoots and roots of the grass, as well as the prairie soil, researchers concluded that plant biomass acted as a temporary nitrogen and phosphorus sink (Seastedt 1988). One study monitored nutrient cycling in restored wetlands on the coast of Maryland (Whigham et al, 2002). It monitored N, P and biomass levels at twelve restored wetlands over a three year period to determine whether nutrient cycling processes were restored. The study found that on a landscape level, all the wetlands had relatively similar nutrient levels in their shoots and soils (Whigham et al, 2002). This implied that regular nutrient cycling processes were established quickly after restoration of the 53 wetlands. These results are important, because they could motivate restoration of degraded prairie wetlands. Bird Habitat Support Grassland bird populations have declined significantly in North America in the past decades. This is primarily due to loss or degradation of grassland habitat, such as prairies, as well as wintering sites, such as coastal wetlands (Baldwin et al, 2007). Efforts to understand the relationship between bird use and grassland characteristics have been made, but more research specific to the Texas coast is needed. Volkert (1992) conducted a study in Wisconsin that investigated the effect of restoring a 200 acre prairie on grassland bird populations. The site was seeded with a mixture of big bluestem, little bluestem, Indian grass, and switchgrass, and the site was monitored over an eight year period. It was found that the number of birds using the site increased almost five-­‐fold from 1986 to 1991. Researchers attribute this increase to the presence of vegetation, which provides nesting cover for grassland birds. However, it was also noted that some species of birds preferred shorter vegetation, while others relied on dense vegetation cover. The implications of this are that within restored prairies, a variety of vegetation structures should be maintained to promote species diversity at the site (Volkert 1992). Similar conclusions were drawn from a study that looked at grassland bird population patterns in a Texas coastal prairie (Baldwin et al 2007). This study investigated the effects of controlled burning on bird use. Researchers found that different bird species favored differing levels of vegetation and shrub densities on the prairie. They suggest that controlled burning should occur in a rotation, leaving some areas with longer periods of growth and other areas where the vegetation is kept sparse (Baldwin et al, 2007). http://blog.nature.org/science/files/2013/02/grasslan
d-­‐birds.jpg http://media.jsonline.com/images/395288
47_Freriks%202a.jpg F IGURE 19: A SPARROW AND A FLYCATCHER ARE TWO COMMON BIRDS TO SEE ON A PRAIRIE GRASSLAND Though the literature on prairie bird support specific to the Texas Coast is not extensive, one study conducted at the Sam Houston National Forest investigated the link between prairie habitat and bird use extensively. Rudolph et al (2014) conducted winter bird surveys starting in 2008 to assess the success of twenty-­‐three prairie restoration projects. These restoration efforts involved the removal of encroaching woody vegetation, controlled fires, and planting of prairie vegetation. Bird surveys were conducted yearly on each prairie and over 30 different species of grassland bird were detected on these 54 prairies over the monitoring years (Rudolph et al 2014). Grassland sparrow populations in particular increased dramatically post-­‐restoration. In prairie sites that had severe woody vegetation encroachment, restoration of isolated prairie grassland patched within the forest increased sparrow populations almost 11-­‐fold over three years. Similar to how the Columbia Bottomlands promote nature tourism through its pristine woodland bird habitat, coastal prairies bring birders in from all areas of the country to view their unique assemblage of species. Prairie habitat is both visually appealing with throngs of flowering perennial plants and sweeping grasses and is a benefit for local communities as the birds that reside in the prairie keep insect populations down around the area and had additional aesthetic beauty with their presence. References Baldwin, H. Q., Grace, J. B., Barrow Jr., W. C., & Rohwer, F. C. (2007). Habitat Relationships of Birds Overwintering in a Managed Coastal Prairie. The Wilson Journal of Ornithology, Vol 119, No. 2, 189-­‐197. Dugas, W. A., Heuer, M. L., & Mayeux, H. S. (1999). Carbon dioxide fluxes over Bermuda grass, native prairie, and sorghum. Agricultural and Forest Meteorology, Vol 93, 121-­‐139. Enwright, N., Forbes, M. G., Doyle, R. D., Hunter, B., & Forbes, W. (2011). Using Geographic Information Systems (GIS) to Inventory Coastal Prairie Wetlands Along the Upper Gulf Coast, Texas. Wetlands, Vol 31, 687-­‐697. Forbes, M. G., Back, J., Doyle, R. D. (2012). Nutrient Transformation and Retention by Coastal Prairie Wetlands, Upper Gulf Coast, Texas. Wetlands, Vol 32, 705-­‐715. Forbes, M. G., & Doyle, R. D. (2012). Tiny Giants: Texas Coastal Prairie Wetlands Are Nutrient Transformers. National Wetlands Newsletter, Vol 34, No 6, 24-­‐27. Ham, J. M., & Knapp, A. K. (1998). Fluxes of CO2, water vapor, and energy from a prairie ecosystem during the seasonal transition from carbon sink to carbon source. Agricultural and Forest Meteorology, Vol 89, 1-­‐14. Kim, J., & Verma, S. B. (1990). Carbon dioxide budget in temperate grassland ecosystem. Boundary-­‐Layer Meteorology, Vol 52, 135-­‐149. Potter, K. N., Torbert H. A., Johnson, H. B., & Tischler, C. R. (1999). Carbon Storage After Long-­‐Term Grass Establishment on Degraded Soils. Soil Science, Vol 164, No 10, 718-­‐725. Risser, P. G., Parton, W. J. (1982). Ecosystem Analysis of the Tallgrass Prairie: Nitrogen Cycle. Ecology, Vol 63, No. 5, 1342-­‐1351. 55 Rudolph, D. C., Plair, D. E., Jones, D., Williamson, H., Shackelford, C. E., Schaefer, R. R., & Pierce, J. B. (2014). Restoration and Winter Avian use of Isolated Prairies in Eastern Texas. Southeastern Naturalist, Vol 13, 52-­‐63. Seastedt, T. R. (1988). Mass, Nitrogen, and Phosphorus Dynamics in Foliage and Root Detritus of Tallgrass Prairie. Ecology, Vol 69, No 1, 59-­‐65. Sims, P. L., & Bradford, J. A. (2001). Carbon dioxide fluxes in a southern plains prairie. Agricultural and Forest Meteorology, Vol 109, 117-­‐134. Suyker, A. E., & Verma, S. B. (2001). Year-­‐round observations of the net ecosystem exchange of carbon dioxide in a native tallgrass prairie. Global Change Biology, Vol 7, 279-­‐289. Tate, C. M. (1990). Patterns and Controls of Nitrogen in Tallgrass Prairie Streams. Ecology, Vol 71, No. 5, 2007-­‐2018. Volkert, W. K. (1992). Response of Grassland Birds to a Large-­‐Scale Prairie Planting Project. The Passenger Pigeon, Vol 54, No. 3, 191-­‐196. Wedlin, D. A., Joost, R. E., & Roberts, C. A. (1996). Nutrient Cycling in Grasslands: An Ecologists Perspective. Nutrient Cycling in Forage Systems, 29-­‐44. Whigham, D., Pittek, M., Hofmockel, K. H., Jordan, T., & Pepin, A. L. (2002). Biomass and Nutrient Dynamics in Restored Wetlands on the Outer Coastal Plain of Maryland, USA. Wetlands, Vol 22, No. 3, 562-­‐574. 56