The scientific literature abounds with research findings indicating how various plants with a high N content decompose faster than plants with a low N content (Bodker et al. 2015). Nitrogen-enrichment enhances plant nitrogen content and accelerates decomposition plant material, including these marshes (Bodker et al. 2015). Phosphorus additions to various marsh plants results in, in the few cases investigated, less live root biomass production (Darby and Turner 2008) and lower soil strength (Turner 2011). The introduced partially-treated sewage has both nitrogen and phosphorous, and particularly nitrogen as ammonium which is 95% of the inorganic nitrogen in the effluent (unpublished results, RET). Kadlec and Wallace’s (2009) encyclopedic review of treatment wetlands schematically summarized the effects of sewage on belowground biomass—natural wetlands have more root biomass that is deeper in the soil than in wetlands receiving sewage (Figs. 3.3 and 3.4 of that book), which is clearly an effect important to understand its consequences to the almost neutrally-buoyant organic soils in this study. The weakened soil strength, therefore, is a function not only of the quantity of belowground biomass, but also the quality of that biomass.
The enhanced organic decomposition and loss of root matter seems to be most significant at the bottom of the rooting depth, which is about 60 cm in this example. That is about the depth of weakened salt marsh soils exposed to the 2010 Deepwater Horizon oil spill (McClenachan et al. 2013) and also from experimental nutrient additions to coastal marshes (Turner 2011; Turner et al. 2009). Soil strength measurements in shallower depths may or may not reveal a response to nutrient additions in experiments (<50 cm; Graham and Mendelssohn 2015; 0–20 cm, Hanson et al. 2016) and the soil matrix may make a difference (Kearney et al. 2015). This weakening of organic soils may not be obvious to the observer looking at only the aboveground vegetation, but can be a pervasive and long-lasting influence on marshes when the buoyant upward forces exceed the rooting strength.
Experiments examining how root biomass respond to nutrients may not capture significant factors when only biomass changes are considered. The larger roots act as metal tie-rods in the ecological cement to give structural strength to the organic whole. We agree with Mayence and Hester (2010) who suggested that “Quantifying the relationship between root length, diameter and tensile strength would benefit our understanding of the interaction between nutrient level and mat strength and is therefore warranted.” and have work underway to investigate this aspect further. The soil strength measured with the shear vane represents a composited force of many roots, whereas the mat breaks away in small intervals as one or a few roots at a time. A change in the size distribution of roots to smaller ones involves a shift in the total soil strength. We did not measure root biomass by size classes, but Hanson et al. (2016) did for a 4 month mesocosm experiment with Spartina alterniflora. They found that the nutrient-enriched experimental units had a reduced number of coarse roots and rhizomes compared to the control units. These results support the conclusion that nutrient-poor soils encourage a larger belowground architectural framework to forage for nutrients and create stronger soils. Hanson et al. (2016) also found that the enriched sites had more finer roots that the unenriched sites, but carefully point out that fine roots decompose more rapidly than coarse roots or rhizomes (Morris et al. 2013; Wigand et al. 2014).
The buoyancy force under flooding is also a function of the force resulting from the submergence of aboveground tissue, which would be increased by the unmeasured upward force of trapped air bubbles (e.g., methane) within the plant’s aerenchyma and in submerged mat, and perhaps released at higher rates with a drop in air pressure (Tokida et al. 2007). The bubble formation is more likely with warmer temperature and more nutrients. The plant’s porous aerenchyma tissue expands as a result of nutrient enrichment adding more buoyancy that is also increased as more of the plant’s basal tissues become submerged from the constant discharge of partially-treated sewage water. The combined force produces uplift on the soil block in which the plant community is anchored. A eutrophic marsh, for example, would more likely become a floating mat than an oligotrophic one (assuming all other factors equal). These rough estimations point to the uncertainty of the size of a surely larger upward pull on the mat—but, there is enough uplifting force to exceed the anchoring soil strength, hence its flotation 5 cm above the water.
Hogg and Wein (1988a) concluded that most of the buoyancy of floating Typha mats was comprised of bubble entrapment, which fluctuated seasonally, and not to tissue development which was about 10–20% of the buoyancy (Hogg and Wein 1988a). The 40–60 cm thick floating mats became detached from their substrate when water level was raised (Hogg and Wein 1988a). The increase in above-ground tissue with the addition of a limiting nutrient (nitrogen in their case) is, therefore, an added factor contributing to mat uplift. The production of gas would be enhanced by decomposition, which was temperature dependent (Hogg and Wein 1988b). There may also be a decreased density in roots, which remains an unmeasured factor.
The bulk density of mineral matter in coastal salt marshes is 2.4–2.5 g cm−3 (Callaway et al. 1997; Craft et al. 1993) and is directly related to soil bulk density, whereas the bulk density of organic matter is around 1.14–1.34 g cm−3 in salt marshes and less than 0.1 g cm−3 in bogs (Krüger et al. 2015). Therefore, the soil mineral content is inversely related to its buoyancy—the higher the organic content, then the greater the buoyancy. This difference between mineral and organic density implies that organic soils may be more susceptible to enhanced nutrient-enriched buoyant uplift forces.
The reduction in soil strength occurs in the apparent absence of nutria grazing, and is in agreement with the opinion of the Louisiana Department of Wildlife and Fisheries’s expert on nutria who was formally asked to make a determination whether nutria were the cause of the loss in this area. He concluded: “Based on biological review of available information, the Department does not consider nutria herbivory to be the primary cause of marsh degradation in the project area” and, they stated that they had never seen such rapid conversion solely due to nutria herbivory (http://www.saveourlake.org/coastal-resources-HammondWorkshop.php#hammondworkshop).
Consideration should be given to the possibility that the partially-treated sewage can vary in toxicity strength and that toxic effluent spikes could be fatal or produce a profound stress to the trees and various species of marsh vegetation. The majority of cypress trees planted with protective collars (5000) within the marsh receiving partially-treated sewage either died, floated out of their anchorage, lodged over or manifested signs of abnormal growth (hypertrophy and stunted height). Some cypress trees planted in the firm soil of the pipeline embankment grew well, but other species on this spoil embankment died after the project began (Bodker et al. 2015).
The nutrient loading rate for this site is within the range for other sites whose soils are described as succumbing to soil strength losses. We can best make these comparisons using the cumulative nutrient loading, and include the synergistic effects of nitrogen and phosphorus additions (Darby and Turner 2008; Turner 2011). But annual estimates give a basis for comparison. Hunter et al. (2009) calculated the loading rates at Joyce WMA using the area of 4047 ha at 2.10 g N m2 year−1, which was equivalent to 21 kg N ha year−1. But the impact area where land turned to water is much smaller (122 ha) which equates to an annual total N loading rate of about 697 kg N ha year−1 over the impacted area. The loading rate would be even higher closer to the first exposure to effluent additions. A 10 ha exposure zone at the beginning of waste delivery, for example, would be 8499 kg N ha year−1. Furthermore, the average concentration of total nitrogen reported by Hunter et al. (2009) was 16.90 mg N L−1, whereas we measured individual values 2 and 3 times higher than this value (unpublished).
These loading rates at the Joyce WMA marsh are within field conditions found elsewhere. The disintegrating marshes of the 5260 ha Jamaica Bay, NY, estuary have some of the highest total nitrogen loading rates at 1096 kg N ha year−1 (Benotti et al. 2007). The annual nitrogen loading for Deegan et al. (2012) long-term nutrient enriched salt marsh site was about 600 kg N ha year−1, and the consequences to the marsh loss became apparent three years later. The nitrogen load at the Caernarvon Mississippi River diversion for the entire flowpath area was 20 kg N ha year−1 (Hyfield et al. 2008), but the first 10% of the receiving area receives a load of 200 kg N ha year−1, and the impacts were not expressed until 19 years later during a high water event (Kearney et al. 2011). The impacts at the site we studied were apparent within one year of effluent application. One implication, then, is that the impacts might appear sooner at sites with higher loading rates or with longer periods of loading.
The total effect of the effluent loading continues to have a consequence to the carbon budget of the area. The 122 ha area converted to open water is about 100 cm deep, and so the amount of carbon released from the marsh is about 450 thousand mt C year−1. There are other issues, including that the un-managed distribution of pathogens into a nutrient- and organic-rich system is a fertile growth medium for organisms potentially compromising the health of people and avians (Anza et al. 2014; Coyner et al. 2003; Spalding et al. 1993; Petrie et al. 2016), and that contaminates local streams—a subject that can be explored in more detail elsewhere. The consequences potentially apply to other marshes with significant vertical water movements and eutrophic conditions.
Wetland-to-open water conversion through the buoyant uplift and the subsequent movement of floating mats has a consequence to interpreting results from experiments using small exclosures used to experimentally test for herbivore grazing effects. Exclosures keep out the herbivore grazers, but also trap and maintain floating organic matter, perhaps to re-connect to the bottom layer. If the mat rises when disconnecting during flooding water, but cannot float away because of restraint by the exclosure wall, then the continuing presence of emergent vegetation could be interpreted as evidence for herbivore grazing outside the plot, whereas none happened. This problem of mis-interpretation is one arising from omitting a disturbed control as part of an experiment. A disturbed control allows access for herbivores, yet maintains the support offered by the wall structure, thereby testing for a ‘cage effect’ on emergent vegetation stability. We have seen exclosure cages at our study area that have one cage wall collapsed, but whose vegetation inside was intact; the area around it was devoid of emergent vegetation. In this case we concluded that herbivore grazing was insignificant.
We conclude that the well-being of this marsh was fatally compromised from the bottom-up influence of the nutrient enrichment. The decline in anchor strength, subsequent mat floatation, and mat decomposition is one way that marsh converts to open water. There are other factors involved, including plant community changes. Untangling the various other effects in a quantified manner can be complicated, and may require a variety of experiments, and many long-term observations (Vermaat et al. 2016).