Estuaries

This note reports an unknown trophic interaction between a mammalian herbivore, the capybara (Hydrochaeris hydrochaeris), and the seagrassRuppia maritima (wigeongrass) and compares the feeding behavior of capybaras to other seagrass grazers. Observations were made in Spring 2002 in the Barra Grande, a small, shallow, moderately stratified, bar-built estuary at Ilha Grande, Rio de Janeiro State, southeast Brazil. The activities of the capybaras were investigated and grazing impacts were quantified in situ. The capybaras were observed feeding onR. maritima during the day and aquatic feeding alternated with periods of feeding on land.R. maritima was the only submerged aquatic vegetation to be consumed by the capybaras. The feeding activity of the capybaras on wige ongrass consisted of alternately diving down to theR. maritima then surfacing; the capybaras spent a significantly greater amount of time under water. In the area where the capybaras foraged 18.1% of the seagrass meadow showed recent grazing scars. Vegetation of recently and not recently grazed areas were compared. Grazing scars, which were slightly elongated, were not completely devoid ofR. maritima but presented reduced standing crop: canopy height, shoot density, and shoot, rhizome, and root biomass were reduced in grazed areas. The grazing patterns observed in capybaras resembled those previously reported in the sirenia, mammals that include two seagrass-eating species.

Transports of nitrate and suspended solids were measured six times from January 1984 until January 1985 in a small freshwater tidal bayou in south-central Louisiana. The bayou and adjacent marshes are influenced by Atchafalaya River discharges, tides, and coastal weather patterns. Large net ebb-directed water transports occurred in winter, spring, and summer, coincident with high river discharges, indicating riverine dominance. A very small net flood-directed water transport occurred in fall, indicating tidally dominated hydrology. Nitrate and suspended solids transports were net ebb-directed in all seasons, but were two orders of magnitude higher during high river flow. Exports changed as hydrology switched from river dominated to tidally dominated, and as concentrations of materials changed. Comparison of suspended solids and nitrate concentrations in the river and bayou shows that these materials were usually lower in the bayou, indicating retention by the marsh/aquatic system.

In recent years, artificial establishment of Spartina alterniflora marshes has become a common method for mitigating impacts to salt marsh systems. The vegetative component of artificially established salt marshes has been examined in several studies, but relatively little is known about the other aspects of these systems. This study was undertaken to investigate the infaunal community of artificially established salt marshes. Infauna were sampled from pairs of artificially established (AE) salt marshes and nearby natural marshes at six sites along the North Carolina coast. The AE marshes ranged in age from 1 yr to 17 yr. Total infaunal density, density of dominant taxa, and community trophic structure (proportions of subsurface-deposit feeders, surface-deposit and suspension feeders, and carnivores) were compared between the two types of marsh to assess infaunal community development in AE marshes. Overall, the two marsh types had similar component organisms and proportions of trophic groups, but total density and densities within trophic groupings were lower in the AE marshes. Soil organic matter content of the natural marshes was nearly twice that of the AE marshes, and is a possible cause for the higher infaunal densities observed in the natural marshes, Using the same three criteria, comparisons of the natural and AE marshes at each of the six locations revealed varying degrees of similarity. Similarity of each AE marsh to its natural marsh control appeared to be influenced by differences in environmental factors between locations more than by AE marsh age. Functional infaunal habitat convergence of an AE marsh with a natural marsh somewhere within its biogeographical region is probable, but success in duplicating the infaunal community of a particular natural marsh is contingent upon the developmental age of the natural marsh and the presence and interaction, of site-specific factors.

Dissolved oxygen concentration below 5 mg 1−1 has characterized the lower tidal portion of the San Joaquin River downstream of Stockton, California, during the summer and fall for the past four decades. Intensive field research in 2000 and 2001 indicated low dissolved oxygen concentration was restricted to the first 14 km of the river, which was deepened to 12 m for shipping, downstream of Stockton. The persistent low dissolved oxygen concentration in the shipping channel was not caused by physical stratification that prevented aeration from vertical mixing or respiration associated wigh high phytoplankton biomass. The low dissolved oxygen concentration was primarily caused bynitrification that produced up to 81% of the total oxygen demand. Stepwise multiple regression analysis isolated dissolved ammonia concentration and carbonaceous oxygen demand as the water quality variables most closely associated with the variation in oxygen demand. Between these two sources, dissolved ammonia concentration accounted for 60% of the total variation in oxygen demand compared with a maximum of 30% for carbonceous oxygen demand. The Stockton wastewater treatment plant and nonpoint sources upstream were direct sources of dissolved ammonia in the channel. A large portion of the dissolved ammonia in the channel was also produced by oxidation of the organic nitrogen load from upstream. The phytoplankton biomass load from upstream primarily produced the carbonaceous oxygen demand. Mass balance models suggested the relative contribution of the wastewater and nonpoint upstream load to the ammonia concentration in the shipping channel at various residence times was dependent on the cumulative effect of ammonification, composition of the upstream load, and net downstream transport of the daily load.

The residence and flushing times of an estuary are two different concepts that are often confused. Flushing time is the time required for the freshwater inflow to equal the amount of freshwater originally present in the estuary. It is specific to freshwater (or materials dissolved in it) and represents the transit time through the entire system (e.g., from head of tide to the mouth). Residence time is the average time particles take to escape the estuary. It can be calculated for any type of material and will vary depending on the starting location of the material. In the literature, the term residence time is often used to refer to the average freshwater transit time and is calculated as such. Freshwater transit time is a more precise term for a type of residence time (that of freshwater, starting from the head of the estuary), whereas residence time is a more general term that must be clarified by specifying the material and starting distribution. We explored these two mixing time scales in the context of the Altmaha River estuary, Georgia, and present a comparison of techniques for their calculation (fraction of freshwater models and variations of box models). Segmented tidal prism models, another common approach, have data requirements similar to other models but can be cumbersome to implement properly. Freshwater transit time estimates from simple steady-state box models were virtually, identical to flushing times for four river-flow cases, as long as boxes were scaled appropriately to river flow, and residence time estimates from different box models were also in good agreement. Mixing time estimates from box models, were incorrect when boxes were imporperly scaled. Mixing time scales vary nonlinearly with river flow, so characterizing the range as well as the mean or median is important for a thorough understanding of the potential for within-estuary processing. We are now developing an imporved box model that will allow the calculation of a variety of mixing time scales using simulations with daily variable river discharge.

A transient network model is applied to the Chesapeake Bay and its tributary estuaries. Calibration of the model is based on only three external parameters: a friction factor that is spatially described, and two global constants required to calibrate a dynamic dispersion relationship that depends on both the local salinity gradient and hydraulic conditions. The transient hydrodynamics and the transient salinity distribution of the Bay and its tributary estuaries are simulated for the period of one month and comparisons made between calculated and observed salinities.

Macrobenthic communities from estuaries throughout the northern Gulf of Mexico were studied to assess the influence of sediment contaminants and natural environmental factors on macrobenthic community trophic structure. Community trophic data were also used to evaluate whether results from laboratory sediment toxicity tests were effective indicators of site-specific differences in benthic trophic structure. A multiple regression model consisting of five composite factors (principal components) was used to distinguish the effects of sediment contaminants and environmental variables on benthic community trophic structure. This model explained 33.5% of the variation in macrobenthic trophic diversity (p<0.001), a variable derived from the distribution of taxas among nine original trophic categories. A significant negative relatinship was found between principal components reflecting concentrations of sediment contaminants and macrobenthic trophic diversity. Detritivores including surface deposit-feeders (SDF), subsurface deposit-feeders (SSDF), and filter feeders (FF) were numerically dominant at 201 random sites, each group accounting for 25–30% of total macrobenthic abundance. The relative abundance of SDFs was considerably lower (12.1±2.9% to 17.1±4.4%) at sites where sediment contaminant concentrations exceeded minimum biological effects thresholds (ER-L values from Long and Morgan 1990 than at sites sampled at random (29.3±5.7%). SSDFs were proportionally more abundant at contaminated sites (42.0±7.7% to 63.6±10.3%) versus random sites (27.5±5.7%), and the relative abundance of SSDFs was positively correlated with concentrations of particular contaminants. Benthic trophic structure was also found to be a function of salinity, where the proportion of SSDFs was negatively correlated with salinity (p=0.035, r=−0.223, n=326). Silt-clay content loaded fairly strongly on the first principal component, but trophic structure parameters were not significantly correlated with sediment grain size or dissolved oxygen (perhaps due, in part, to covariation). Results from laboratory sediment toxicity tests with mysids were predictive of differences in macrobenthic trophic structure in situ (i.e., mysid survival was negatively correlated with %SSDF; p<0.001, r=−0.292, n=326). Results from laboratory sediment toxicity tests with ampeliscid amphipods were not indicative of site-specific differences in benthic trophic structure. Results from this study demonstrated that sediment contaminants can be quite important in structuring macrobenthic communities in soft-bottom estuarine habitats. The fact that macrobenthic trophic diversity decreased significantly with increasing sediment contamination indicates that important general differences in benthic community function may exist between contaminated and random sites. These data suggest that benthic trophic structure analysis may be an effective tool for assessing integrated community responses to chronic sublethal exposure and may be useful for assessing toxicological responses at ecologically relevant levels of organization.

Distribution patterns of suspended sediments and sea surface temperatures in, Mobile Bay were derived from algorithms using digital data from the visible, near infrared, and infrared channels of the Advanced Very High Resolution Radiometer (AVHRR) on the NOAA-TIROS-N satellite. Closely spaced AVHRR scenes for January 20, 24, and 29, 1982, were compared with available environmental information taken during the same period. A complex interaction between river discharge, winds, and astronomical tides controlled the distribution patterns of suspended sediments. These same variables, coupled with air temperatures, also governed the distribution patterns of sea surface temperatures.