Chapter 23
GLOBAL CHANGE AND TIDAL FRESHWATER WETLANDS: SCENARIOS AND IMPACTS Scott C. Neubauer & Christopher B. Craft
This chapter was originally published in the book „Tidal Freshwater Wetlands“. The copy attached is provided by Margraf Publishers GmbH for the author‘s benefit and for the benefit of the author‘s institution, for non-commercial research, and educational use. All other uses, reproduction and distribution are prohibited.
Tidal Freshwater Wetlands, edited by Aat Barendregt, Dennis Whigham & Andrew Baldwin 2009, viii + 320pp.; (incl. 16 colour plates), 21 x 29,7 cm, hardbound ISBN 978-3-8236-1551-4 © Copyright 2009, Margraf Publishers GmbH
Backhuys Publishers, Leiden Margraf Publishers, Weikersheim, 2009
TABLE OF CONTENTS PART 1: INTRODUCTION AND HISTORICAL USE 1. Tidal freshwater wetlands – an introduction to the ecosystem Andrew H. Baldwin, Aat Barendregt & Dennis F. Whigham 2. Human activities in European tidal freshwater wetlands Ies S. Zonneveld & Aat Barendregt 3. Human uses of tidal freshwater wetlands on the USA east coast Erik Kiviat
PART 2: ECOSYSTEM DESCRIPTION 4. Hydrogeomorphology and sedimentation in tidal freshwater wetlands Gregory B. Pasternack 5. Plant communities of tidal freshwater wetlands of the continental USA and southeastern Canada Mary Allessio Leck, Andrew H. Baldwin, V. Thomas Parker, Lisa Schile & Dennis F. Whigham 6. Plant communities of European tidal freshwater wetlands Eric Struyf, Sander Jacobs, Patrick Meire, Kai Jensen & Aat Barendregt 7. Animal communities in North American tidal freshwater wetlands Christopher W. Swarth & Erik Kiviat 8. Animal communities in European tidal freshwater wetlands Aat Barendregt, Tom Ysebaert & Wim J. Wolff 9. Invasive plants in tidal freshwater wetlands of the USA east coast Erik Kiviat
PART 3: PROCESSES 10. Primary production in tidal freshwater wetlands Dennis F. Whigham 11. Characteristic aspects of the tidal freshwater zone that affect aquatic primary production Stefan Van Damme, Eric Struyf, Tom Maris, Tom Cox & Patrick Meire 12. Carbon flows, nutrient cycling, and food webs in tidal freshwater wetlands Stuart E.G. Findlay, William C. Nieder & Serena Ciparis
PART 4: CASE STUDIES 13. Northeastern North American case studies – New Jersey and New England Mary Allessio Leck & Caitlin M. Crain 14. Tidal freshwater wetlands of the mid-Atlantic and southeastern United States James E. Perry, Donna M. Bilkovic, Kirk J. Havens & Carl H. Hershner 15. Tidal freshwater wetlands of the Mississippi River deltas Charles E. Sasser, James G. Gosselink, Guerry O. Holm Jr. & Jenneke M. Visser 16. Tidal freshwater wetlands of Alaska Jonathan V. Hall 17. Ecological consequences of a change in tidal amplitude in tidal freshwater wetlands Aat Barendregt, Peter Glöer & Frank Saris 18. Water and nutrient balances of the experimental site Mariapolder, The Netherlands Wladimir Bleuten, Wiebe Borren, Esther Kleinveld, Lieke B. Oomes & Tiemo Timmermann
PART 5: RESTORATION, CONSERVATION, AND FUTURE DEVELOPMENTS 19. Restoration of tidal freshwater wetlands in North America Andrew H. Baldwin 20. Restoration of European tidal freshwater wetlands Aat Barendregt 21. Conservation of tidal freshwater wetlands in North America Dennis F. Whigham, Andrew H. Baldwin & Christopher W. Swarth 22. Conservation of tidal freshwater wetlands in Europe Ericia Van den Bergh, Annick Garniel, Roger K.A. Morris & Aat Barendregt 23. Global change and tidal freshwater wetlands: scenarios and impacts Scott C. Neubauer & Christopher B. Craft 24. Synthesis and perspectives for the future Dennis F. Whigham, Aat Barendregt & Andrew H. Baldwin
REFERENCES INDEX LATIN NAMES INDEX KEYWORDS
© Copyright 2009 Backhuys Publishers, Leiden, The Netherlands Backhuys Publishers is a division of Margraf Publishers GmbH Scientific Books, Weikersheim, Germany. All rights reserved. No part of this book may be translated or reproduced in any form by print, photoprint, microfilm, or any other means without prior written permission of the publisher. Margraf Publishers GmbH Scientific books, P.O. Box 1205, D-97985 Weikersheim, Germany.
Chapter 23
GLOBAL CHANGE AND TIDAL FRESHWATER WETLANDS: SCENARIOS AND IMPACTS Scott C. Neubauer * & Christopher B. Craft
* corresponding author - e-mail:
[email protected]
Abstract: Tidal freshwater wetlands (TFW) have developed in a gradually yet constantly changing environment, but current and predicted future rates of environmental change, driven in large part by human activities, are often faster than those in the recent geological past. Global warming and increases in atmospheric CO2, sea level rise and salt water intrusion, and eutrophication-related changes in water quality are likely to affect the distribution, floral/faunal composition, biogeochemical functioning, and persistence of TFW. The current distribution of TFW indicates that these wetlands can persist across broad gradients of climate, hydrology, and nutrient loading. Comparisons across these gradients provide insight into the changes that may occur as individual wetlands are subjected to a changing environment. Because salt water intrusion is going to change one of the defining features of TFW (i.e., the presence of fresh water), this stressor is likely to have the largest long-term impact. In systems where river discharge is insufficient to offset salt water intrusion, TFW are likely to migrate upstream if space is not limited by topography or human development. Community changes are likely in current TFW as saltintolerant species are replaced by brackish species. Salt water can also affect nutrient sorption and availability, and may alter rates of vertical accretion and carbon/nutrient sequestration. It is likely that TFW will be impacted by multiple, simultaneous environmental changes. Understanding feedbacks between these stressors and TFW responses is necessary to accurately forecast how global changes will impact these ecologically unique ecosystems in a future climate. Plant nomenclature follows USDA Plants Database (http://plants.usda.gov). Zoological nomenclature follows Integrated Taxonomic Information System (http://www.itis.gov/) Keywords: environmental change, climate change, anthropogenic disturbance, global warming, elevated CO2, sea level rise, salt water intrusion, eutrophication, water quality
INTRODUCTION In their unique position at the border between terrestrial and aquatic ecosystems, where rivers become estuaries, tidal freshwater wetlands (TFW) are sensitive to a wide range of environmental disturbances that are forecast for the coming decades and centuries. Global and regional climate change will affect TFW via upstream effects (e.g., changes in watershed precipitation, runoff, nutrient inputs), downstream effects (e.g., changes in tidal flooding and salt exposure due to sea level rise), and atmospheric changes (e.g., increased temperature and greenhouse gases). Further, local landscape modifications (e.g., shoreline hardening, changes in land use) will directly influence coastal wetlands and feed back to magnify or moderate additional climate changes. At some level, these disturbances are linked, with interactions
between the mechanisms underlying environmental change and the responses of TFW themselves (Fig. 1). The title of this chapter includes the phrase “global change”, but factors that affect TFW ecosystems can vary on regional to local levels. For example, current predictions suggest that higher latitudes will warm at a greater rate than subtropical and tropical regions (Meehl et al. 2007), and there will be differences in the degree of warming within single continents (Christensen et al. 2007). Similarly, sea level rise will not proceed at the same rate in all TFW, which are located across multiple continents and a multitude of individual river systems, each with its own geomorphology, tectonic setting (i.e., uplifting vs. subsiding), and local-scale impacts (e.g., local subsidence due to groundwater withdrawal for industrial use; Tangley 1988). Each of these factors will affect rates of relative sea level rise, making it difficult to forecast
Tidal Freshwater Wetlands, pp. 253-266 Edited by A. Barendregt, D.F. Whigham, A.H. Baldwin; © 2009 Backhuys Publishers, Leiden, The Netherlands
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future trajectories across broad geographic areas. Thus, this chapter will provide a generalized description of how TFW might respond in our changing environment. We focus this chapter on four global environmental changes: global warming and increases in atmospheric CO2, sea level rise, salt water intrusion (linked with sea level rise, but discussed separately), and eutrophication-related changes in water quality. There are other environmental issues that may impact TFW – chemical pollution/contamination, invasive/exotic species, changes in storm frequency/intensity, and atmospheric deposition of NOx and SO42-, to name a few – but we feel that the four impacts we have highlighted will be the most significant and applicable across the entire geographic range where tidal freshwater wetlands occur.
Global change scenarios An ensemble of 21 global climate models was used in the 2007 reports by the Intergovernmental Panel on Climate Change (IPCC) as a basis for global and regional-scale forecasts of future climate and rates of sea level rise (Christensen et al. 2007, Meehl et al. 2007). Those projections are briefly discussed here, with the caveat that sub-continent to continent-scale projections can disguise a large amount of smaller scale (local) variability, especially with respect to precipitation (Good & Lowe 2006). A doubling of atmospheric CO2 concentrations is expected to increase the global mean surface temperature by 3 °C (“likely” range of 2 to 4.5 °C of
warming; Meehl et al. 2007). On the sub-continent scale, warming will be similar in eastern North America (median model output = +3.2 °C for 2080-2099, relative to 19801999), southern Europe (+3.5 °C) and northern Europe (+3.2 °C; Christensen et al. 2007). Within eastern North America, annual warming will be greater in the New England region than in the southeast or Gulf of Mexico coasts. In Europe, winter warming will be greatest in northeastern Europe, whereas summer warming will be greater in the southern part of the continent. In both eastern North America and northern Europe (north of 48 °N), precipitation is projected to increase by 7 and 9%, respectively, with greater relative precipitation increases in the northern portions of these subcontinental regions. In contrast, the ensemble model forecast is for a 12% decrease in annual precipitation in southern Europe (Christensen et al. 2007). The balance between precipitation and evapotranspiration, which is likely to increase in a warmer climate, will affect freshwater availability and river discharge into the tidal freshwater zone. Rates of global eustatic sea level rise have averaged ~1.8 mm/yr since the 1960s and are projected to increase to 3.8 mm/yr during the last decade of the twenty-first century (based on the SRES A1B scenario of future population growth and greenhouse gas emissions; Meehl et al. 2007). Due to subsidence, rates of relative sea level rise exceed those of eustatic sea level rise along the US Atlantic coast and in Europe along the North and Baltic Seas (excluding Scandinavia where uplift is occurring; Emery 1980). Furthermore, there is considerable spatial variability in rates of relative sea
Figure 1. Environmental and anthropogenic stressors operate on local to global scales and impact on tidal freshwater wetland community structure, ecological functioning, and ecosystem persistence. Arrows indicate driving forces and feedbacks.
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level rise across regional scales (Titus & Narayanan 1995, CPSL 2001), making it difficult to forecast rates of relative sea level rise for specific regions and river systems. For example, sea level rise along the US Atlantic coast, as inferred from tide gauges, ranged from 1.8 to 4.3 mm/yr during the twentieth century (geometric mean = 2.9 mm/yr, n = 20 sites) with rates near 10 mm/yr along the Gulf of Mexico (Titus & Narayanan 1995). Future projections of cultural eutrophication in tidal freshwater zones, and estuaries and coastal waters in general, are highly dependent on population growth and management scenarios (Seitzinger et al. 2002). Decisions regarding land use, water management, and best management practices could lead to either increased or decreased future nutrient loads relative to current conditions. For example, the Patuxent River (Maryland) and Choptank River (Maryland, Delaware) have N and P inputs that are 4 to 20 times greater at present than during the 1700s due to wastewater nutrient inputs (Patuxent) and agriculture (Choptank) (Fisher et al. 2006). Despite recent improvements in N and P removal during wastewater treatment in the Patuxent, increases in the volume of wastewater discharge due to increased population have resulted in no net change in nutrient loads to the estuary. In contrast, water quality in the tidal Scheldt estuary (Belgium, The Netherlands) has improved since the mid1970s, with significant reductions in dissolved phosphorus, inorganic nitrogen, and silica due to upgraded sewage treatment and restrictions on waste disposal (Soetaert et al. 2006). The sensitivity of water quality to management decisions is in sharp contrast to climate change and sea level rise where, due to time lags and feedbacks in the climate system, future warming and sea level rise will still occur even if greenhouse gas emissions are stabilized (Meehl et al. 2007).
Anthropogenic disturbances Historically, the upper ends of estuaries, where TFW are located, were preferred locations for development due to the presence of fresh water (i.e., rivers) for drinking and agriculture, and access to oceanic shipping routes via estuaries (see: Chapters 2, 13, 17). Although the focus of this chapter is on environmental impacts on TFW, a brief discussion of anthropogenic forcing is warranted as there are significant linkages and feedbacks between human actions, environmental disturbances, and TFW responses (Fig. 1). Globally, approximately 60% of the population lives within 100 km of the coast (Vitousek et al. 1997), and this percentage is expected to increase in the future, resulting in even more pressures on coastal wetlands and freshwater resources. Localized impacts associated with human settlement can directly result in the loss of TFW (e.g., via filling or draining of wetlands), affect tidal water quality (e.g., via conversion of forest to agriculture, or incomplete wastewater treatment), or negatively impact the ability of TFW to migrate upstream and inland (i.e., via shoreline development and hardening). Watershed land use changes such as deforestation may increase the availabil-
ity of mineral sediments and lead to higher rates of vertical wetland accretion, a potential benefit from the perspective of dealing with sea level rise. On broader regional and global scales, fossil fuel combustion and land use changes (primarily deforestation) are largely responsible for rapid increases in atmospheric greenhouse gas concentrations (e.g., CO2, CH4, N2O), resulting in changes in climate and increasing rates of sea level rise (Denman et al. 2007).
GLOBAL WARMING AND GREENHOUSE GASES Plant communities Rising temperatures have been implicated in changing distributions of plants in both terrestrial and wetland habitats (e.g., Harris & Cropper 1992, Iverson & Prasad 1998, Kittel et al. 2000). Changes in the latitudinal distribution of TFW species will depend on changes in temperature, the availability of suitable habitat (in terms of hydroperiod, substrate type, salinity), and the dispersal ability of each species. Many species that are common in TFW have a wide tolerance to environmental conditions (including temperature) as evidenced by their wide geographic distribution (Table 1). Thus, global warming, by itself, is unlikely to result in broad shifts in TFW vegetation. For species whose southern limit is determined by tolerance to high temperatures, global warming may result in a displacement of those species northward without a net expansion of their range. Given the high species diversity of many TFW and redundancy within the community (in terms of plant characteristics and habitat preferences), shifts in the distribution of individual species probably will have few impacts on ecosystem functioning. Increasing CO2 levels may result in an increasing dominance of C3 vs. C4 plants in TFW due to the different responses of these plant types to elevated CO2 (e.g., Poorter 1993)
Animal communities The geographic distribution of animal species and rates of metabolic activity are strongly influenced by temperature. In aquatic environments, higher temperatures decrease dissolved O2 levels by reducing the amount of O2 that the water can hold (i.e., lower O2 saturation level) and by increasing rates of O2 consumption by heterotrophic processes. In shallow TFW creeks, which can already go hypoxic or anoxic during summer low tides (e.g., Neubauer & Anderson 2003), decreased O2 availability may further limit utilization of tidal creeks. Temperature also plays a key role in growth and survival of different fish species. For example, the year-class strength of Micropogonias undulates (Atlantic croaker), an estuarine/marine species that sometimes feeds in tidal freshwater regions, is positively correlated with warm winter temperatures and the over-winter survival of juveniles,
Chapter 23 – Global change and tidal freshwater wetlands: scenarios and impacts
Table 1. Geographic ranges of selected TFW species along the Atlantic coast of the United States. While recognizing that TFW occur elsewhere in North America and globally, the Atlantic coast of the USA is 256 used as an example because it presents a large climatic gradient across a >10° latitudinal S.C. Neubauer & C.B. Craft range (Brunswick, Georgia: mean annual temperature = 21 °C vs. Acadia, Maine = 7 °C; National Climatic Data Center, http://www.ncdc.noaa.gov). Species could potentially migrate northward in response to Table 1. Geographic ranges of selected TFW species along the Atlantic coast of the United States. While recognizing that TFW occur elseglobal warming. Frost-free days represent the minimum “growing season” length for successful growth of where in North America and globally, the Atlantic coast of the USA is used as an example because it presents a large climatic gradient across species. in parentheses indicate the species hasvs.been found in =those is notData Center1). a >10°each latitudinal rangeStates (Brunswick, Georgia: mean annualthat temperature = 21 °C Acadia, Maine 7 °C;states, Nationalbut Climatic common in the coastal counties.inData are from thewarming. U.S. Department of represent Agriculture’s Plants“growing Database Species could potentially migrate northward response to global Frost-free days the minimum season” length for successful growth of each species. States in parentheses indicate that the species has been found in those states, but is not common in the (http://plants.usda.gov). coastal counties. Data are from the U.S. Department of Agriculture’s Plants Database2.
������� ������ ������� (sweetflag) ������� �������� (northern wild rice) ����� ��������� (broadleaf cattail) ������� ��������� (rice cutgrass) ��������� ��������� (arrow arum) ��������� ���������� (arrowleaf tearthumb) ������ �������� (crowned beggarticks) �������� ���������� (rosemallow) �������� ��������� (baldcypress) ����� �������� (water tupelo) rather than the larval supply (Hare & Able 2007). In contrast, anadromous Morone saxatilis (striped bass) have higher larval survival following cold winters (Lippson et al. 1979); this may be due to high spring runoff that fuels phytoplankton growth and subsequent zooplankton densities, thus providing a food source for the larvae. Many of the mammals that utilize TFW have broad geographic ranges and therefore are unlikely to be dramatically affected by gradually rising temperatures. As a notable exception, Myocastor coypus (nutria) is sensitive to freezing temperatures (Willner et al. 1979, as cited in Odum et al. 1984) and may therefore be able to shift its distribution northward as the climate warms. Since reptiles and amphibians are ectotherms, the geographic distribution of these animals is determined by local temperature regimes. Many herpetofauna species have either their northern or southern distribution limit within the Chesapeake Bay region (Musick 1972, as reported in Odum et al. 1984). With increasing temperatures, these animals may expand their distributions northward if suitable habitat is available, with a parallel northward movement of the southern end of their range if high temperatures there exceed the thermal tolerance of the animals. For the many birds that utilize TFW, the effects of global warming are likely to be indirect and related to changes in plant and prey food resources (Najjar et al. 2000). In the case of migratory birds, climate change effects in the wintering or breeding grounds could have large impacts on the abundances of birds that temporarily use TFW as feeding areas (see: Chapter 7).
���������� ����� ���� 90 95 100 110 110 110 145 160 160 231
���������� ���������� �������� �������� ME NC (SC, GA) ME GA ME FL ME SC (GA) NH (ME) FL ME SC (GA) MA SC (GA) MA GA (FL) MD/DE FL NC (VA) SC (GA)
tal community (plant + soil) CO2 effluxes (Q10 = 1.9 to 3.2; Neubauer et al. 2000). Thus, marsh-to-air CO2 fluxes due to plant and soil respiration are likely to increase by ~21 to 79%, given a 3 to 5 °C increase in temperature (e.g., as predicted by 2100 for the mid-Atlantic USA region; Najjar et al. 2000). Similarly, losses of dissolved inorganic carbon to tidal waters are also likely to increase as water temperatures warm (Q10 = 1.6 to 2.0; Neubauer & Anderson 2003). Rates of methanogenesis also are strongly affected by temperature, with reported Q10 values for wetlands averaging 4.1 (Segers 1998). Net CH4 effluxes are driven by the balance between rates of methanogenesis and methane oxidation. By analogy with data showing that rates of SO42- reduction at higher temperatures are limited by rates of hydrolysis and fermentation (Weston & Joye 2005), warming may increasingly lead to methanogenesis being limited by the supply of low molecular weight precursors such as acetate. This, however, may be complicated by changes in the relative importance of hydrogenotrophic vs. acetotrophic methanogenesis with changing temperatures. At higher temperatures, rates of acetotrophic methanogenesis typically decrease; this may be a function of changing substrate availability (Megonigal et al. 2004). In a tidal freshwater forest soil, methane oxidation was CH4limited, implying that the efficiency of CH4 oxidation in wetland forests is not likely to change in response to global warming (Megonigal & Schlesinger 2002). However, given that methane oxidation is generally less sensitive to temperature (average Q10 = 1.9; Segers 1998) than methanogenesis and because tidal marshes typically occupy lower (wetter) 15 landscape positions than tidal forests, warming may increase net CH4 fluxes from non-forested TFW. Together, increasBiogeochemistry ing rates of organic matter decomposition and losses of soil C as CO2 and CH4 imply that organic matter accumulation Temperature is the master variable driving most metaand preservation rates will be lower in a future, warmer clibolic processes, so global warming is likely to increase rates mate, unless increased respiration is offset by increased inof both autotrophic and heterotrophic processes within TFW. puts through primary production or allochthonous sources. In TFW, temperature response values for plant canopy resThese increased greenhouse gas emissions from TFW can piration (Q10 = 2.0, Nietch 2000) are similar to those for to-
1 2
http://www.ncdc.noaa.gov http://plants.usda.gov
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feed back to further increase rates of global warming and sea level rise, although this effect may be minimal since the area of TFW is small globally. Plant production is likely to increase with global warming and elevated atmospheric CO2 levels (e.g., Bazzaz 1990, Drake et al. 1996), although factors such as limited nutrient availability and increased salinity may reduce the CO2 fertilization effect. A meta-analysis of elevated CO2 studies suggested that wetlands are less responsive to elevated CO2 with respect to C accumulation in plant and soil pools than are upland systems (Luo et al. 2006). To our knowledge, no field studies in TFW have examined productivity or net C and N accumulation under elevated CO2 so here and elsewhere in this chapter we use studies of brackish/salt marshes to indicate probable directions of TFW responses to climate change stressors. In a long-running (18 year) study in a brackish marsh in Maryland, elevated CO2 stimulated the growth of a C3 sedge (Schoenoplectus americanus, three square) but not a C4 grass (Spartina patens, saltmeadow hay) (Erickson et al. 2007). Further, the effect of elevated CO2 was greatest in rainy, low salinity years. An interaction between rising sea levels and elevated CO2 has resulted in an increase in S. americanus biomass and a decline in that of S. patens in areas where these two species co-occur. By extension, the effects of elevated CO2 in diverse TFW ecosystems are likely to depend on a combination of factors including species composition, salinity, and sea level. Several laboratory studies have looked at the growth of individual TFW species under elevated CO2. Photosynthetic rates increased in Taxodium distichum (baldcypress) and Orontium aquaticum (goldenclub) under elevated CO2, regardless of whether the soil surface was flooded or not (Megonigal et al. 2005). However, the biomass response of the plants differed by flooding regime. T. distichum showed increased biomass under elevated CO2 in the non-flooded treatment; in contrast O. aquaticum increased biomass in both flooding treatments. The differing responses likely reflect the natural landcape positions occupied by these two species – T. distichum grows in drier environments and is more sensitive to flooding than O. aquaticum. Similarly, Acer rubrum (red maple) seedlings, which are less tolerant to flooding than T. distichum, did not show a biomass response to elevated CO2 when grown in saturated or flooded soils (Vann & Megonigal 2002). It remains to be determined if the responses of wetland trees to elevated CO2 differ between the weeks to months time scale of elevated CO2 laboratory experiments and the decades to centuries life span of the trees. For T. distichum and O. aquaticum, there were strong positive correlations between whole-plant photosynthesis and biomass with CH4 emissions, implying a tight coupling between plant processes (e.g., exudation, detrital inputs) and microbial processes such as methanogenesis (Megonigal & Schlesinger 1997, Vann & Megonigal 2003). In contrast, Whiting and Chanton (1996) reported that CH4 emissions from Typha latifolia (broadleaf cattail) did not vary as a function of atmospheric CO2 levels. They suggested that the primary factor controlling CH4 fluxes through T. latifolia was
the rate at which soil CH4 could enter the roots, not rates of primary production or the stomatal conductance of the plant itself. Across a variety of freshwater wetland types, there is a positive correlation between net primary production and CH4 emissions (Whiting & Chanton 1993).
SEA LEVEL RISE Tidal freshwater wetlands are extremely vulnerable to rising sea level through increasing inundation and salt water intrusion (Rheinhardt & Hershner 1992). Accelerated sea level rise resulting from global warming is projected to lead to wetland loss through submergence (Park et al. 1989, Brinson et al. 1995, Moorhead & Brinson 1995) and habitat conversion as saline wetlands migrate landward (Park et al. 1991). Global warming also is projected to increase interannual variability of precipitation, leading to greater frequency of drought and floods (Karl et al. 1995, Mahlman 1997) and greater variability in freshwater discharge of rivers and streams. Reduced freshwater discharge may interact synergistically with rising sea level, leading to rapid loss and habitat conversion of TFW. This section focuses on the effects of changing hydroperiod (frequency/duration of flooding) on TFW, although there is some unavoidable overlap with salt water intrusion since these disturbances are often going to occur in parallel. A more complete discussion of potential salinity effects on TFW can be found in the Salt water intrusion section.
Ecosystem migration Simulation modelling of accelerated sea level rise along the southeastern United States coast suggests that TFW will be severely reduced during the next 100 years. For example, using the Sea Level Affects Marshes Model (SLAMM) (Park et al. 1986), we estimate that, along the 160 km Georgia coastline, tidal swamp forest coverage will be reduced by 97 km2 (-24%) during the next 100 years whereas tidal freshwater marsh areas will remain unchanged (Table 2). The large reduction in tidal swamp forest area is due to topographic limitations to upstream and horizontal wetland migration in the upper reaches of the estuaries. A simple, elevation-based model for the state of Delaware predicted that a sea level rise of 61 cm by 2,095 would flood 91 km2 of land, resulting in the loss of 21% of Delaware’s wetlands (Najjar et al. 2000). However, that model did not consider the ability of the wetlands to grow vertically (by accumulating mineral or organic material) or to migrate horizontally (along the Delaware coast, there is little development to restrict lateral wetland movement), so the projected wetland loss probably represents an upper estimate for the state. Along the northeastern USA coast and in the upper tidal reaches of many systems, the loss of TFW will be even greater because more abrupt topographic relief and higher levels of watershed development will limit inland migration of TFW, and a narrowing
Chapter 23 – Global change and tidal freshwater wetlands: scenarios and impacts
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Table 2. Predicted change in area (in km2) of selected (wet)land cover types along the southeastern (Georgia) USA coast in response to the SRES A1B sea level rise scenario as modeled by SLAMM. The 2 Table 2. Predicted change areausing (in km ) of selected types along the southeastern (Georgia) coast response to the simulation wasinrun the A1B mean(wet)land scenario,cover which assumes a 52 cm increase in SLRUSA during theinnext SRES A1B100 sea level scenario as modeled SLAMM. The simulation was run using the A1B mean scenario, which assumes a 52 cm years.rise Source: Craft et al. (inbypress). increase in SLR during the next 100 years. Source: Craft et al. (2009).
Dry land Non-tidal swamp forest Tidal swamp forest Freshwater marsh Tidal freshwater marsh Brackish marsh Salt marsh Tidal flat Estuarine open water
���� ���� ���� ����� 5,008 1,838 413 64 79 417 1,116 11 742
���� ���� ���� ����� 4,385 2,089 316 65 80 458 890 26 1,091
������� ������ -12 +14 -24 +2 +1 +10 -20 +136 +47
rise, the flooding regime will not change and therefore plant productivity should not be affected. If sea level rise rates exceed vertical accretion, plant productivity will increase until the wetland is at its optimal elevation relative to sea level; beyond this point, productivity is likely to decrease rapidly. The optimum marsh elevation and depth of inundation will vary regionally depending on tide range and other factors Plant communities (McKee & Patrick 1988, Morris et al. 2002). The effects of flooding on TFW primary production are complicated by TFW vegetation is vulnerable to inundation and salt water the fact that species composition typically varies along an intrusion associated with sea level rise although the effects elevation gradient (Odum et al. 1984, Mitsch & Gosselink vary depending on the species. In a mesocosm experiment, 1993). Thus when the productivity of one species declines flooding depths of +5 cm and +20 cm significantly reduced due to excessive flooding, there may be a shift in plant comaboveground biomass of Sagittaria lancifolia (bulltongue munity composition to more flood-tolerant plants, with no arrowhead) and Spartina patens (Spalding & Hester 2007). change in net productivity. Shifts in salinity associated with Aboveground biomass of Panicum hemitomon (maidenrising sea levels are also likely to affect plant biodiversity cane), however, increased with flooding depth. Other studies and productivity. have similarly shown that flooding increases P. hemitomon aboveground biomass (Willis & Hester 2004) or at least has Flooding not only affects the growth of established TFW no negative effect (Koch & Mendelssohn 1989, Pezeshki et plants, but also seed germination and the growth of seedlings. al. 1991). The effects of inundation and elevated salinity (up Field and greenhouse studies with natural seed banks from to salinity = 6) reduced biomass production more than eitidal freshwater wetlands have demonstrated that biodiversither treatment singly (Spalding & Hester 2007). On the other ty and seedling emergence decrease as the depth of flooding hand, tidal freshwater forests show negative effects of floodincreases (Baldwin et al. 2001, Peterson & Baldwin 2004b). ing (both depth and duration), with the interactive effects of For example, field experiments in the Patuxent River, Maryincreased flooding plus salinity exceeding those of flooding land, showed that lowering wetland sods by 10 cm decreased or salinity alone (Allen et al. 1996, Conner et al. 1997). Difspecies richness of emerging seeds by 26%, whereas raising ferent plant types (i.e., C3 vs. C4) may respond differently to the sods by 10 cm increased species richness by 42% (Baldchanging water levels (Saunders et al. 2006) and plants that win et al. 2001). Additionally, growth (as total stem length) are near the lower end of their range in the tidal frame will was ~two times greater in the raised treatment than in the be most sensitive to increases in flooding depth. more flooded plots (Baldwin et al. 2001). Further, seedling emergence decreased with the addition of sediments, which There is evidence that, in saline Spartina alterniflora is a likely consequence of increased flooding (Peterson & (smooth cordgrass) coastal wetlands, increasing inundation Baldwin 2004b). Thus, increased flooding due to sea level leads to increased net primary production, up to a point. Morris et al. (2002) showed that for S. alterniflora in South 16 rise is likely to negatively impact biodiversity in TFW, especially with respect to annuals since the germination of annual Carolina, Net Primary Production (NPP) reaches a maximum seeds is inhibited by flooding, while perennials can persist at depths of 40 to 60 cm below mean high tide. At greater via vegetative growth even if flooding increases (Baldwin flooding depths, marsh vegetation is abruptly replaced by et al. 2001). unvegetated tidal mud flats. If this pattern holds in TFW, the response of the vegetation community to sea level rise will be linked to rates of vertical wetland growth. If the marsh is growing vertically at a rate similar to relative sea level channel will reduce the upstream area suitable for TFW formation. The effects of accelerated sea level on tidal freshwater wetlands in other regions and on other continents also will vary depending on the geomorphology of each region.
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Sediment deposition and vertical accretion Feedbacks between hydroperiod, sedimentation, plant productivity/species composition, and organic matter production/decomposition will ultimately influence wetland accretion and nutrient accumulation rates. In TFW, both sediment deposition (see: Chapter 4, Pasternack & Brush 1998, Neubauer et al. 2002, Darke & Megonigal 2003, Morse et al. 2004) and longer-term sediment accretion rates (e.g., Merrill 1999, Merrill & Cornwell 2000) are typically greater in sites that are flooded more frequently and for a longer duration, or on levees immediately adjacent to tidal creeks. Thus, sediment deposition and accretion rates are likely to increase as sea level rises, unless TFW can grow vertically
at a rate approximating sea level rise. Recent rates of TFW accretion (median rate = 0.76 cm/yr, based primarily on 137Cs and 210Pb dating) are generally greater than rates of sea level rise (Neubauer 2008), suggesting that most TFW are wellpoised to deal with accelerating rates of sea level rise. If rates of soil organic matter decomposition increase due to rising temperatures and salt water intrusion, TFW may become further dependent on allochthonous sediments for vertical growth. This is potentially a problem for TFW in sedimentpoor systems (e.g., northeast United States) and suggests that wetlands in sediment-rich watersheds will be more resilient to higher rates of sea level rise (Morris et al. 2002).
SALT WATER INTRUSION The time frame until salt water intrusion will affect TFW will vary between riverine systems depending on site-specific rates of sea level rise and changes in freshwater riverine discharge. Future changes in watershed precipitation and runoff may accelerate rates of salt water intrusion (i.e., with decreased discharge), slow upstream movement of the salt front (with slight to moderate increases in discharge), or even result in a down-estuary movement of the salt front (with large increases in discharge). The effects of changing discharge on the salinity of TFW can already be observed over seasonal to annual scales. For example, in the Potomac River (Fig. 2), the salinity of monitoring station TF2.2 (TF = tidal freshwater) was always ≤ 0.1, except during the record low flows of summer 1999 when the salinity at this site increased to only 0.3. At a second tidal freshwater station (TF2.4, ~25 km downstream of TF2.2), surface water salinity ranged from 1-3 during low-flow summer months, but was ≤ 0.1 during most of the year. During high discharges (e.g., most of 1996), the salt front is pushed downstream to the point that freshwater conditions can dominate in normally brackish waters (e.g., station RET2.1, ~15 km downstream of TF2.4; RET = river-estuarine transition).
Plant communities Figure 2. Potomac River discharge and salinity in tidal fresh and oligohaline waters. Shaded region represents 25th and 75th percentiles of monthly discharge for 1930 to 2006 (monthly medians denoted by curve in middle of shading). Solid discharge curve shows actual discharge for each month from 1995 to 1999. Asterisks indicate when the highest or lowest discharges for a particular month during the 77 years of record occurred during the interval shown on this figure. The salinities at stations RET2.1 (most seaward; dotted line) and TF2.4 (solid gray line) are more sensitive to river discharge than at station TF2.2 (most landward; solid black line). Discharge data are collected in the Potomac River at the Little Falls Pump station (station # 01646500) by the United States Geological Survey (USGS)3. Salinity data are collected approximately biweekly as part of the Chesapeake Bay monitoring program4.
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http://waterdata.usgs.gov http://www.chesapeakebay.net
The high plant diversity of TFW is, in part, due to the lack of salt stress in these ecosystems and it is clear that salinity plays a key role in structuring tidal wetland plant assemblages (e.g., Odum et al. 1984, Mitsch & Gosselink 1993). The role of salinity may be due to the inability of freshwater plants to deal with the increased ionic strength or with metabolites associated with saline waters (e.g., H2S; Koch et al. 1990). Multiple studies have reported that species diversity along estuarine transects often is greater in TFW than in brackish wetlands (e.g., Anderson et al. 1968, Latham et al. 1994, Perry & Atkinson 1997), but this is not always so (S. Pennings pers. comm.). Thus, with salt water intrusion, vascular plant diversity is likely to decline, with a concomitant shift in species composition, as freshwater
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species are replaced with more salt-tolerant species. In the southeastern USA, net primary productivity may increase since aboveground biomass is greater in brackish wetlands than TFW (S. Pennings pers. comm.). Relatively few studies have repeatedly surveyed the vegetation of TFW over long time scales (decades) to determine if environmental changes are affecting plant assemblages. Following construction of a tide gate in the Savannah River estuary (Georgia and South Carolina), the salt front moved upstream 3 to 10 km and caused a displacement of freshwater plant species by brackish ones (Pearlstine et al. 1993). Perry and Hershner (1999) found that the relative abundance of oligohaline species such as Spartina cynosuroides (big cordgrass) and Carex hyalinolepis (shoreline sedge) increased in the oligohaline portion of a TFW on the Pamunkey River, Virginia between 1974 and 1987 and suggested that salt water intrusion may have played a role. Despite some changes in species composition, there was not a significant change in plant diversity. In the Delaware River, there has been a long-term (since 1800) net loss of TFW vegetation and an upstream movement of S. alterniflora (Schuyler et al. 1993). Just as long-term changes in salinity at the freshwater end of estuaries can lead to vegetation changes, the lack of a directional shift can reflect the absence of a long-term trend in water salinity. Analysis of aerial images dating back to 1953 showed that there was not an upstream shift in wetland vegetation in two Georgia estuaries, perhaps because highly variable river discharges masked the slight increases in salinity that would be expected from a rising sea level (Higinbotham et al. 2004). In TFW, which are typically composed of a mixture of annual and perennial species, the perennial species may be the best indicators of long-term environmental change since they tend to persist once established, unless the environment changes significantly (Warren & Niering 1993). Annual species, in contrast, are generally opportunistic so changes in their abundance and distribution may be more reflective of short-term fluctuations (e.g., a regional drought) rather than long-term climate changes (Perry & Hershner 1999). In contrast with tidal marshes, where vegetation changes typically result in the replacement of one herbaceous species assemblage with another, salt water intrusion into tidal freshwater swamps can lead to a dramatic state change that converts tree-dominated swamps into herbaceous-dominated wetlands. Because Taxodium distichum, Nyssa aquatica (water tupelo), and Fraxinus pennsylvanica (green ash) are all sensitive to salinity (Conner et al. 1997), salt water intrusion is resulting in the death of large areas of freshwater wetland forests throughout the southeastern USA (DeLaune et al. 1987, Pezeshki et al. 1990, Allen 1992, Krauss et al. 2000). This swamp-to-marsh transition in response to sea level rise is opposite the normal marsh-to-swamp successional pattern often seen at the freshwater end of tidal estuaries (Conner et al. 2004, Tufford 2005).
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Animal communities Spatial surveys of species abundance in estuarine waters have often identified a species minimum (the Artenminimum) between salinities of 5 and 8 (e.g., Deaton & Greenberg 1986). Thus, salt water intrusion may result in a decrease in biodiversity of nekton in the upper tidal reaches of estuaries. As the salt front moves inland, estuarine nekton are likely to follow and true freshwater species may disappear from the (formerly) tidal freshwater zone. Anadromous fishes that are often found in the tidal freshwater zone and its fringing wetlands may be relatively unaffected since the critical zone that determines the survival of larval and juvenile anadromous fishes can extend to salinities as high as 10 (Dovel 1981). In the future, oligohaline wetlands may serve the same nursery roles currently filled by TFW. As saline wetlands migrate landward, predation on forage fish will increase as marine predators also migrate (Deegan 2002). However, consumer stress models predict that prey items will be less inhibited by stress (e.g., salinity) than consumers, resulting in increased prey abundances and increased competition with increasing stress (Menge & Olson 1990). These models may not apply if the stress results in shifts in species composition. Losses of preferred food sources can lead to the loss of consumers, regardless of whether the consumer itself is directly affected by salt water intrusion. For example, Eudocimus albus (white ibis) nestlings require freshwater food sources (Bildstein et al. 1990); salt water intrusion into freshwater wetlands can negatively impact Procambarus spp. (crayfish) populations and cause the dispersal of white ibis colonies (Michener et al. 1997).
Biogeochemistry Salt water intrusion is likely to affect rates and pathways of catabolic metabolism, nutrient cycling and ecosystem exchanges, and the carbon sequestration potential of TFW. These biogeochemical changes will be driven by several mechanisms: 1) physio-chemical effects due to changes in ionic strength and pH; 2) microbial reactions and competitive interactions due to increased sulfate availability; and 3) suppression and/or inhibition of biological processes due to changes in ionic strength, pH, and H2S concentrations. Additionally, biogeochemical processes can be affected by salinity-induced changes in plant community composition (as discussed above), which can impact rates of root O2 loss as well as the amount and lability of organic matter added to soils. Following strict thermodynamic theory, competition between soil microbes for electron donors (e.g., organic carbon, H2) results in a sequence of terminal electron accepting reactions that proceed in order from aerobic respiration to NO3- reduction, Mn(IV) reduction, Fe(III) reduction, SO42reduction, and methanogenesis (Schlesinger 1997). If sufficient electron donors are available, as one electron acceptor is exhausted (e.g., O2) the next most energetically favorable
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electron acceptor (e.g., NO3-) will be utilized until it too is depleted, and so on. However, due to spatial heterogeneity with respect to electron acceptor and electron donor distributions, multiple processes will often occur within the same soil volume. Because SO42- reduction rates in freshwater wetlands are often limited by the availability of SO42-, salt water intrusion and increased SO42- delivery should increase SO42- reduction rates, at the expense of methanogenesis. This competition between SO42- reducers and methanogens is a possible explanation for the observation that CH4 production and emission rates typically increase toward the head (freshwater end) of estuaries (e.g., Bartlett et al. 1987, Neubauer et al. 2005b). Microbial competition may also explain decreases in CH4 vs. CO2 emissions from tidal freshwater sediments exposed to a 10 d laboratory salinity intrusion event (Weston et al. 2006). Furthermore, the addition of salt water to these sediments increased total metabolism by a factor of ~two (as measured by total CO2 + CH4 generation). Similarly, Nietch (2000) found a trend of increasing metabolism (CO2 + CH4 emissions) in tidal freshwater marsh cores exposed to brackish water (salinity = 19). In a laboratory experiment using unvegetated whole cores from a Delaware River tidal freshwater marsh, the addition of brackish water (salinity = 5) stimulated both CO2 and CH4 emissions relative to freshwater-flooded cores over a several month period (N. Weston pers. comm.). Changes in TFW biogeochemistry can feed back to increase rates of global warming (via increased CO2 and possibly CH4 emissions) or moderate climate change (via possibly depressed CH4 emissions). Salt water impacts on soil metabolism and plant productivity may also affect sediment deposition, vertical accretion, and organic matter accumulation in TFW. Indeed, across estuarine gradients, rates of marsh accretion and C, N, and P accumulation are similar in tidal freshwater and brackish wetlands, but lower in salt marshes (Fig. 3; Craft 2007). In part, the differences in short-term sediment deposition reflect differing proximity to upland watershed sedi-
a)
ment sources (Odum 1988). The combination of increased rates of soil metabolism, greater rates of root decomposition in salt marshes (Craft 2007), and reduced plant productivity suggests that salt water intrusion will have negative consequences with respect to the ability of TFW to accrete organic matter and grow vertically with rising sea levels. Spalding and Hester (2007) grew Panicum hemitomon in mesocosms exposed to fresh or brackish water (salinity = 0, 2, 4, or 6) for an entire growing season. In the salt-exposed mesocosms, plant productivity was typically lower than in the freshwater mesocosms; this difference in productivity was used as an explanation for the lower soil organic matter content in the brackish mesocosms. In a natural TFW, the effects on soil organic matter accumulation (and therefore vertical accretion) may not be as dramatic as in single species mesocosms. As plant species composition changes in response to salt water intrusion, the accumulation of organic matter from salt-tolerant species may offset that lost due to salt stresses on freshwater wetland plants. Furthermore, some wetland plants that can survive in oligohaline and mesohaline waters (e.g., Spartina cynosuroides) are more recalcitrant than strictly freshwater plants (e.g., Dunn 1978), especially the broad-leafed species such as Peltandra virginica (arrow arum), Pontederia cordata (pickerel weed), and Nuphar lutea (spatterdock), so a higher fraction of the plant biomass may be incorporated into the soil profile. However, in TFW swamps, salt water intrusion will result in the replacement of relatively recalcitrant tree biomass (C/N ratio ~ 35 to >40; e.g., Schlesinger 1978, Gomez & Day 1982) with more labile herbaceous marsh biomass (C/N ratio ~ 16 to >25; e.g., Dunn 1978, Neubauer et al. 2005a). If salinity also accelerates rates of organic matter decomposition, the transition from swamps to herb-dominated tidal wetlands may have significant consequences with respect to the ability of the ecosystem to accrete sufficient organic matter for vertical growth. Overall, the evidence suggests that replacement of TFW with salt marshes likely will lead to reduced carbon
b)
Figure 3. (a) Short-term (6 month) sedimentation and long-term (40 year) vertical accretion and (b) organic C, N, and P accumulation in soils of tidal freshwater, brackish and salt marshes of the Altamaha River, Georgia, USA. Sedimentation rates were determined using feldspar marker layers (n=3 per marsh). Accretion and accumulation rates are calculated from 137Cs dating of soil cores (n=2 per marsh). Data replotted from Craft (2007).
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sequestration and retention of nitrogen, which is especially important for estuaries that typically are nitrogen-limited and thus susceptible to nitrogen enrichment and eutrophication (see: Water quality and eutrophication, below). Many of the reactions involved in TFW nitrogen cycles are potentially sensitive to salt water intrusion. The elevated ionic strength associated with salt water can lead to increased, although perhaps transient, soil-to-water NH4+ fluxes as salt water cations (e.g. Na+ and Mg2+) displace NH4+ that was bound to soil particles (Seitzinger et al. 1991). Increased rates of organic matter mineralization are also likely to increase NH4+ availability. Modelling efforts for a freshwater lake (Haringvliet, The Netherlands) that will be restored to tidal estuarine conditions indicate that NH4+ efflux from the sediments will increase by ~2.5-fold soon after salinity restoration but decline to a value only ~15% greater than freshwater conditions after 6 months (Canavan et al. 2007a). This decline occurs as a new NH4+-sediment equilibrium is reached following salinity intrusion and as the availability of labile organic matter decreases. With increases in porewater NH4+, there may be a net suppression of N2 fixation since nitrogenase can be inhibited by free NH4+ (e.g., Howarth et al. 1988). The modelling efforts of Canavan et al. (2007a) indicated that knowing how the partitioning between denitrification and dissimilatory NO3- reduction to NH4+ (DNRA) responds to salinity is a key parameter in understanding how NO3- cycling in freshwater sediments will respond to salt water intrusion. The cycling and fates of NO3- can also be affected by salt water intrusion due to physiological effects of salinity on nitrifying and denitrifying microbes (e.g., MacFarlane & Hebert 1984, Furumai et al. 1988, Stehr et al. 1995) and H2S inhibition of denitrification (Brunet & GarciaGil 1996, An & Gardner 2002). Across the estuarine gradient, DNRA is generally more important than denitrification in brackish and marine sediments (Tobias et al. 2001); thus salt water intrusion may increase N retention (i.e., DNRA) within TFW ecosystems. Similarly, Rysgaard et al. (1999) found that a simulated salt water intrusion event into freshwater sediments resulted in pronounced decreases in denitrification, nitrification, and coupled nitrification-denitrification. The effects of salt water intrusion on TFW nitrogen cycling are likely to be lowest in systems that are near the freshwater-oligohaline transition, where wetlands may historically have been exposed to elevated salinities during low discharge years (Fig. 2) and may support halotolerant microbial populations (e.g., Magalhães et al. 2005). Salt water intrusion can also affect the cycling of phosphorus in TFW soils. In a survey of tidal wetlands along the salinity gradient in the Cooper River estuary, South Carolina, Sundareshwar and Morris (1999) found that exchangeable PO43- concentrations were significantly lower in a brackish marsh (salinity = 16.5) than in a tidal freshwater/oligohaline site (salinity = 0.4). In part, these differences reflect in situ “extraction” of soil PO43- that occurs in more saline waters (i.e., salts limit the accumulation of PO43- on soil surfaces). However, other factors including mineral surface area and
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soil mineralogy play a key role in soil P dynamics. In TFW soils that are exposed to O2 (via root O2 loss or atmospheric inputs during drainage), the formation of iron oxyhydroxide minerals can trap significant amounts of PO43-. This “iron curtain” (Chambers & Odum 1990) can be important in reducing wetland-to-water PO43- fluxes from creekbank soils. With the addition of SO42-, however, ferrous (reduced) Fe can be sequestered in Fe-S minerals such as pyrite (FeS2), resulting in increased mobilization of sediment-bound P (Caraco et al. 1989, Lamers et al. 2001). Due to the interactions between Fe and H2S, differences in soil Fe content can influence the availability of H2S, with subsequent implications regarding H2S toxicity to plants and microbial populations. Nietch (2000) compared SO42- and H2S distributions within the Edisto and Cooper Rivers, South Carolina. There was a sulfide deficit (less H2S at similar SO42- concentrations) within the Cooper River wetlands and this trend was associated with higher Fe in the Edisto River wetlands. Salt water intrusion is also likely to impact organic P cycling by affecting rates of plant production (see above). Further, since phosphatase enzymes are used to increase P availability when inorganic P is insufficient to meet plant or microbial demands, increasing PO43- availability due to the breakdown of the iron curtain may result in lower phosphatase activity and correspondingly higher rates of organic P sequestration following salt water intrusion.
WATER QUALITY AND EUTROPHICATION Nutrient enrichment is a problem affecting many estuaries (Ryther & Dunstan 1971, Smith 1998) and, increasingly, tidal wetlands (McClelland & Valiela 1998, Wigand et al. 2003). Most coastal ecosystems are N-limited (Howarth et al. 2002) and N-enrichment promotes eutrophication, hypoxia and anoxia (Mitsch et al. 2001). In the United States, nutrient enrichment also facilitates invasion by non-native species such as Phragmites australis (common reed; Galatowitsch et al. 1999). Little is known, however, about the effects of nutrient enrichment on TFW.
Plant communities The effect of nitrogen fertilization on TFW varies by species. Most in situ fertilization studies with Peltandra virginica have found that additions of N and/or P do not increase plant biomass or the biomass nutrient content (Whigham & Simpson 1978b, Walker 1981, Booth 1989, Chambers & Fourqurean 1991, Morse et al. 2004), although a significant fertilization effect on P. virginica was observed in a mesocosm study (Nietch 2000). Nutrient additions to common TFW plants Spartina cynosuroides (Booth 1989), Panicum hemitomon (DeLaune et al. 1986), and Typha domingensis
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(southern cattail, grown in mesocosms; Nietch 2000) have documented a positive fertilization effect. Long-term additions of nutrient-rich wastewater to a non-tidal stand of Taxodium distichum resulted in significant increases in growth (Hesse et al. 1998). Fertilization of a Zizaniopsis miliacea (giant cutgrass)-dominated TFW in Georgia (USA) revealed that N, rather than P, is the primary limiting nutrient for plant growth (Frost et al. 2009). After two years of nutrient additions, aboveground biomass in the N and N+P plots (2,100 g/m2) was more than double that in control (900 g/m2) and P-treated (1,000 g/m2) plots. Nitrogen:phosphorus ratios of leaves ranged from 14-27 across fertilized and unfertilized plots (Frost et al. 2009), also suggesting N limitation based on a 30:1 N:P ratio (mol:mol) threshold between N versus P limitation (Koerselman & Meuleman 1996). Likewise, other studies of TFW have inferred N limitation of primary production based on N:P ratios in aboveground plant tissue, accumulating sediments, and surface soils (Morse et al. 2004) although, as discussed by Chambers and Fourqurean (1991), it is not always a simple matter to determine which nutrient (if any) limits plant growth in TFW. In contrast to TFW where relatively little work on the effects of nutrient enrichment has been conducted, salt marshes have a rich history of nutrient enrichment studies. From this literature, we can make inferences regarding how nutrient enrichment will affect TFW, especially herb-dominated sites. In a review paper, Deegan (2002) postulated that nutrient enrichment will alter structural complexity of salt marshes through increased plant height but not stem density. Nutrient enrichment also may alter the balance between NPP and decomposition (see: Biogeochemistry, below), leading to subsidence, habitat fragmentation, and marsh loss. Deegan (2002) also suggested that, not surprisingly, tidal wetlands located at the upper, freshwater ends of estuaries will be among the first to be affected by enrichment because of their proximity to nutrient sources. Nutrient enrichment also may lead to changes in plant diversity though this has not been tested in TFW. Crain (2007) added N and P to species-rich oligohaline wetlands in Maine (salinity < 5). After three growing seasons, plots receiving N+P were dominated by Spartina pectinata (prairie cordgrass) and Solidago sempervirens (seaside goldenrod), two species that produce large amounts of aboveground biomass. In the control plots, these species were present, but in low abundance (
