Evolving paradigms and challenges in estuarine and coastal eutrophication dynamics in a culturally and climatically stressed world

Abstract

Coastal watersheds support more than one half of the world’s human population and are experiencing unprecedented urban, agricultural and industrial expansion. The freshwater-marine continua draining these watersheds are impacted increasingly by nutrient inputs and resultant eutrophication, including symptomatic harmful algal blooms, hypoxia, finfish and shellfish kills and loss of higher plant and animal habitat. In addressing nutrient input reductions to stem and reverse eutrophication, phosphorus (P) has received priority traditionally in upstream freshwater regions, while controlling nitrogen (N) inputs has been the focus of management strategies in estuarine and coastal waters. However, freshwater, brackish and full-salinity components of this continuum are connected structurally and functionally. Intensification of human activities has caused imbalances in N and P loading, altering nutrient limitation characteristics and complicating successful eutrophication control along the continuum by reducing only one nutrient. Several examples increasingly indicate the need for dual N and P input constraints as the only effective nutrient management option for long-term control of eutrophication. Climatic changes, specifically increased variability in freshwater discharge associated with more severe storms and intense droughts, interact closely with nutrient inputs to modulate the magnitude and relative proportions of N and P loading. The effects of these interactions on phytoplankton production and composition were examined in two neighboring North Carolina lagoonal estuaries, the New River and Neuse River Estuaries; which are experiencing concurrent eutrophication and climatically-driven hydrologic variability. Efforts aimed at stemming estuarine and coastal eutrophication in these and other similarly-impacted estuarine systems should focus on establishing N and P input thresholds that include effects of hydrologic variability, so that eutrophication and harmful algal blooms can be controlled over a range of current and predicted climate change scenarios.

Introduction

More than half of the Earth’s human population resides in coastal water- and air-sheds (Vitousek et al., 1997; NOAA 2012; Ache et al., 2013). Accelerating agricultural, urban and industrial development in these “sheds” has put unprecedented pressure on the ecological condition and sustainability of downstream riverine, estuarine and coastal waters (Bricker et al., 1999; Boesch et al., 2001; Conley et al., 2009). Nutrient over-enrichment has been identified as a prime cause for water quality and habitat degradation (Nixon 1995; Paerl 1997; Boesch et al., 2001; Elmgren and Larsson 2001; Rabalais and Turner 2001; Diaz and Rosenberg 2008). In addition, hydrologic modifications including upstream reservoir construction, agricultural and urban surface and groundwater water withdrawal, have altered water flow rates and paths, sedimentation rates and optical properties of receiving waters. These modifications strongly affect estuarine and coastal water and habitat quality (Cloern 2001; Boesch et al., 2001; Rabalais and Turner 2001; Humborg et al., 2007). While human modification of coastal water and airsheds directly, and often negatively, impacts water and habitat quality of these systems (Bricker et al., 1999; National Research Council 2000), climatic factors such as warming, more extreme storms, floods and droughts modulate these impacts (Paerl and Huisman 2008, 2009; Jeppesen et al., 2010).

Human and climatically induced stresses strongly interact in estuarine watersheds. These interactions affect the delivery and effects of freshwater discharge and its nutrient load on the activity and composition of microalgal primary producers in geographically diverse estuarine and coastal ecosystems. Here, we evaluate how human alterations of watershed nutrients and hydrology affect coastal eutrophication dynamics and the measures needed to control it and associated habitat degradation in a more crowded, and climatically extreme world. This issue will be addressed in the context of managing the freshwater to coastal continuum, the physiographic and biogeochemical gradient coupling water- and air-sheds to coastal ecosystems (Fig. 1).
Nutrient enrichment and limitation in estuarine and coastal ecosystems: Historical and current perspectives

The dominant nutrient limitation (of primary production) paradigms applied to this continuum for more than one-half a century, were that phosphorus (P) availability controlled primary production in freshwaters, while nitrogen (N) was the dominant limiting nutrient in the more saline downstream estuarine and coastal waters (Ryther and Dunstan 1971; Schindler 1975; Nixon 1995; Boesch et al., 2001; Smith and Schindler 2009). Brackish waters often exhibit sensitivity to both N and P inputs (Fisher et al. 1999; Rudek et al., 1991; Elmgren and Larsson 2001; Paerl and Piehler 2008). Recent analyses of diverse nutrient limitation studies in both freshwater and marine ecosystems indicate that these paradigms may be “eroding” (Elser et al., 2007; Lewis and Wurtsbaugh 2008; Sterner 2008; Conley et al., 2009; Lewis et al., 2011). Increasingly, incidences of N&P “co-limitation”, i.e., the stimulation of primary production by the addition of N and P in combination, where N or P alone stimulate production far less, have been reported (Elser et al., 2007; Lewis et al., 2011). Also, exclusive N limitation in freshwater ecosystems (as opposed to exclusive P limitation) is more common than believed previously (Elser et al., 2007; Lewis et al., 2011). Concurrently, recent estuarine and coastal studies indicated that N&P and/or P limitation are geographically widespread (Peeters and Peperzak 1990; Elmgren and Larsson 2001; Sylvan et al., 2006; Paerl and Justi

2011; Laurent et al., 2012).
Liebig’s “Law of the Minimum”: Theory vs. practice

In examining nutrient limitation of primary producers, Liebig’s Law of the Minimum, also referred to as Liebig's Law or the Law of the Minimum, is an operational principle first developed in agricultural science by Carl Sprengel (1839) and later popularized by Justus von Liebig (cf., Brown 1942). It states that plant growth is controlled not by the total amount of resources available, but by the scarcest resource (limiting factor). From an aquatic plant production perspective, the yield is proportional to the amount of the most limiting resource (i.e., nutrient, light, etc.). If the limiting resource is a nutrient, it follows that if the deficient nutrient is supplied, yields may be improved to the point that some other nutrient is needed in greater quantity than the water body can provide. In the case of N&P co-limitation, the balance between N&P supply may be very near the demand so that the addition of both stimulates primary production, but the addition of either one alone does not.

The application of Liebig’s law has been demonstrated in chemostat cultures with single phytoplankton species, where nutrient supplies can be controlled tightly and growth yield is regulated by an essential nutrient supplied in amounts below those needed to maintain optimal cellular growth. These cultures can be manipulated to show that several nutrients can be supplied simultaneously, at ratios and rates very close to those needed to maintain balanced growth (i.e., “Redfield ratio”, Redfield 1958; Redfield et al., 1963). Only a slight increase in the supply of one nutrient will shift the control on yield to the other nutrient. In natural systems supporting complex phytoplankton communities, such nutrient supply shifts can occur as a result of variability in external nutrient inputs (loads), internal nutrient cycling, and sediment-water column nutrient exchange. Furthermore, changes in plant community composition, due to death, grazing, plant-microbe and higher trophic level interactions (e.g., nutrient regeneration from zooplankton grazers up to fish) can also affect nutrient availability and hence nutrient limitation. It is therefore possible to envision how even subtle shifts in nutrient supply, community composition and biogeochemical cycling can affect the nature and complexity of nutrient limitation.
Nutrient limitation paradigms

Historically, there is a much longer and geographically diverse line of nutrient limitation data available for freshwater than marine ecosystems; most likely because the symptoms of nutrient over-enrichment and eutrophication have been more evident and problematic in freshwater ecosystems, dating back several centuries to the establishment and expansion of agriculture (i.e., rapid increase in chemical fertilizer use), the industrial revolution, and urbanization (Thienemann 1915; Naumann 1921; Parma 1980).

Freshwater and estuarine ecosystems tend to have larger water- and air-shed areas relative to their surfaces than do more open coastal marine ecosystems or large, deep lakes. Water replacement rate, relative to the volume of that water body (i.e., flushing rate), is often high in shallow enclosed systems. Therefore, from a nutrient input and enrichment perspective, these systems are influenced heavily by their water- and air-sheds. They tend to be dominated by N inputs, since N compounds are often soluble and associated with a wide variety of organic and inorganic sources (e.g., plants, soils, atmospheric emissions and combustion products, and microbial transformations), whereas P is associated with rocks and soils, where it is often insoluble and therefore unavailable. Furthermore, N, unlike P, exists in significant gaseous forms, and is more mobile and easily transported and transformed in the geospheres, biospheres and atmospheres. As such, freshwater to oligohaline estuarine systems are often enriched in N relative to P and frequently exhibit P limitation.

Exceptions to this paradigm are often related to specific watershed geochemical characteristics. Silicon (Si) may be deficient in watersheds that are dominated by non-silicious rocks and soils (e.g., carbonates). This pattern can lead to Si limitation, especially for diatoms, in downstream N and P-enriched waters (Justíc et al., 1995; Dortch and Whitledge 1992). Furthermore, construction of upstream dams and reservoirs can promote “trapping” of Si-containing soils and sediments, causing Si deficiency downstream (Humborg et al., 2007; Chai et al., 2009).

Phosphorus limitation occurs commonly in relatively undisturbed water- sheds (cf. Wetzel 2001), supporting the early conclusion that P is the limiting nutrient in most freshwater ecosystems (cf., Likens 1972). Unfortunately, only a few “undisturbed” air- and water-sheds remain to evaluate this paradigm. Today, agricultural, urban and industrial expansions have altered the landscape, and amounts and patterns of nutrient loading to freshwater ecosystems. These activities have increased both N and P loading, with wastewater inputs and runoff from land clearing and the establishment of farmland and urban centers as dominant nutrient sources. The early recognition of P as a primary limiting nutrient in these systems (Likens 1972; Schindler 1975), and the linkage of P loading to freshwater eutrophication (Vollenweider 1968), provided the impetus for focusing on P input reductions (Schindler and Vallentyne 2008). Indeed, such reductions were effective in stemming and reversing problematic symptoms of eutrophication, nuisance algal blooms, food web disruption, bottom water hypoxia, and degradation of planktonic and benthic habitats (Wetzel 2001; Schindler and Vallentyne 2008).

Nutrient loading dynamics have changed dramatically over the past several decades. While P reductions were pursued actively, human population growth and agricultural and urban expansion in watersheds were paralleled by increased rates of N loading (Peierls et al., 1991; Howarth 1998), often exceeding those for P (Rabalais 2002; Galloway and Cowling 2002). In the Baltic Sea region, subjected to several centuries of human nutrint enrichment, effective control of eutrophication requires considering total amounts and ratios of N and P discharged to a nutrient-sensitive, river-fjord-sea continuum (Elmgren and Larsson 2001; Conley et al. 2009). Similarly, single nutrient input reductions, including a P-detergent ban and improved wastewater treatment for P during the 1980’s in North Carolina’s (USA) Neuse River System, helped arrest freshwater blooms, but failure to reduce N inputs simultaneously exacerbated blooms in downstream N-sensitive estuarine waters (Paerl et al., 2004). In both cases, parallel N and P input reductions were required to stem eutrophication (Elmgren and Larsson 2001; Paerl 2009). In Florida’s (USA) extensive lake-river-estuary systems, excessive N loading, much of it from expanding wastewater and agricultural discharges, was identified (in addition to P) as supporting eutrophication (Kratzer et al. 1981). N₂ fixing cyanobacteria often dominate in Lake Okeechobee, Florida’s largest lake. However, non-N₂ fixing genera (e.g. Microcystis), and facultative N₂ fixing genera (e.g. Cylindrospermopsis, Lyngbya) compete effectively for reactive N when it is available and only fix N₂ when other available N is depleted (Moisander et al., 2012). In all cases, both N and P reductions are needed to control eutrophication and harmful (toxic, hypoxia-generating) cyanobacterial bloom genera (Howarth et al., 2000; Havens et al., 2001).

Lake Erie, USA-Canada, seemed to have “recovered” from eutrophication due to P (but not N) reduction programs (Schindler 2012), but eutrophication resurged with the phytoplankton domination by non-N₂ fixing cyanobacteria (Microcystis sp.; Lyngbya sp.). Oligo- to mesohaline regions of large estuaries and coastal bays and seas (e.g., Chesapeake Bay, Albemarle-Pamlico Sound, NC, Florida Bay, FL, Coastal North Sea and Baltic Sea; Moreton Bay, Australia) also reveal N and P co-limitation (Peeters and Peperzak 1990; Rudek et al., 1991; Fisher et al., 1999; Elmgren and Larsson 2001; Watkinson et al., 2005; Ahern et al., 2007); largely because previously-loaded P and N are retained and recycled. This observation is especially true for P, which exists only in soluble (orthophosphate, dissolved organic P) and particulate forms, which are cycled between the water column and bottom sediments (Vollenweider 1968; Boynton and Kemp 1985; Wetzel 2001).

In contrast, N exists in multiple forms, including dissolved (nitrate, nitrite, ammonium, organic N), particulate, and gaseous; P has no significant gaseous forms. Gaseous forms of N (N₂, N₂O, NO, NO₂, NH₃, volatile organic N compounds), are produced by microbial transformations, including ammonification, anammox, denitrification and nitrification (Capone et al., 2008), which can escape into the atmosphere. In particular, denitrification is a major N loss mechanism. However, this process does not keep up with externally-supplied “new” N inputs, especially in systems impacted by N over-enrichment (Seitzinger 1988; Nixon et al., 1996).

Agricultural and domestic synthetic fertilizers, fossil fuel combustion, wastewater treatment, and a wide range of industrial chemical processes are major sources of biologically reactive N to estuarine and coastal waters, and have increased dramatically over the past half-century (Galloway and Cowling 2002; US EPA 2011). In Northern Gulf of Mexico waters receiving discharge from the Mississippi River Basin, agricultural and urban N inputs in the Basin have increased so rapidly that the receiving marine waters now exhibit P limitation during spring with elevated runoff, while N limitation prevails during the drier summer months (Sylvan et al., 2006). Anthropogenically-generated P has also increased, but in many instances not nearly as rapidly as N (Justić et al., 1995), especially in places where P detergent bans and improved wastewater treatment for P have been implemented. In many intensively farmed, urbanized and industrialized regions, however, historic and current P loads are still quite high because of P-saturated soils, and continued P fertilizer applications. As a result, residual P supplies in sediments have remained high and available.

While a N input “glut” is occurring due to expanding anthropogenic inputs, a fraction of the N supplied to water bodies is “lost” as N₂ via denitrification or converted to other gaseous forms (e.g., NH₃, N₂O and NO emissions) (US EPA 2011). N₂ fixation rates are not sufficient to offset N losses via denitrification and anammox in these systems, perpetuating N limited conditions (Paerl and Scott 2010). Thus, despite receiving ever-increasing anthropogenic N inputs, these systems can still assimilate these inputs and become more eutrophic, without becoming exclusively P limited, due to N losses via denitrification and other gas-generating processes, while P is internally cycled. This phenomenon appears widespread in meso- to eutrophic freshwater and marine ecosystems, which exhibit N limitation or N and P co-limitation (Granéli et al. 1999; Elmgren and Larsson 2001; Elser et al., 2007; Sterner 2008; Paerl and Piehler 2008; Finlay et al., 2010; Lewis et al., 2011).

Interestingly, nutrient management efforts in freshwater components of the estuarine continuum continue to focus largely on “P only” reduction strategies (cf. Schindler and Vallentyne 2008; Schindler et al. 2008), despite pioneering studies in the 1960’s showing a role for N in freshwater eutrophication (cf., Goldman 1981; Wetzel 2001) and more recent studies demonstrating sensitivity of a range of lakes and reservoirs to N inputs (Elser et al., 2007; North et al., 2007; Lewis and Wurtsbaugh 2008; Finlay et al., 2010; Xu et al., 2010; Paerl et al., 2011b; Spivak et al., 2011; Lewis et al., 2011). This approach is based on the assumption that cyanobacteria can supply N via nitrogen (N₂) fixation (Schindler et al., 2008). However, at the ecosystem-level, only a fraction, usually far less than 50%, of primary production demands are met by N₂ fixation, even when P supplies are sufficient (Scott et al. 2008; Paerl and Scott 2010). As a result, “perpetual N limitation” can occur in many freshwaters due to seasonal inorganic N drawdown (Scott et al 2009). This chronic N deficit appears to be even more pronounced for estuarine and coastal waters (Howarth et al., 1988). Indications are that N₂ fixation is controlled by factors in addition to just P availability (Paerl 1990; Scott and McCarthy 2010). N-limitation may persist in aquatic ecosystems, even in the presence of N₂ fixers. Therefore, external N inputs play a key role in controlling primary production along the continuum (cf., McCarthy et al., 2007; Conley et al., 2009; Paerl 2009; Paerl et al., 2011b).

From a nutrient management perspective, important questions and research needs include; 1) How do patterns of reactive N drawdown lead to seasonal N limitation or N+P co-limitation? 2) How is this drawdown partitioned between phytoplankton N demand and denitrification? 3) What is the critical balance (i.e., threshold) between N:P and loading rates and removal processes (i.e., uptake, denitrification) over episodic, seasonal, and multi-annual time scales?

The interacting roles of climate change: Warming and increased hydrologic variability

Climatic changes, specifically warming and altered rainfall amounts and patterns, strongly interact with nutrient enrichment in modulating eutrophication dynamics. Temperature controls algal metabolism and growth rates, and the controls appear taxa-specific (cf., Paerl et al., 2011a) (Fig 2). Most notable is the stimulatory effect of warming on growth rates of cyanobacteria, the only prokaryotic phytoplankton group (Paerl and Huisman 2008). Being “bacterial”, cyanobacterial species tend to show growth optima that are in the 25-30 ^oC range, in contrast to eukaryotic taxa, which typically exhibit growth maxima at lower temperatures (Paerl et al., 2011a). Therefore, longer, warmer growing seasons favor cyanobacterial species (Jöhnk et al., 2008; Paerl et al., 2011a). Also, there are taxa-specific effects of warming on eukaryotic phytoplankton groups. For example in the Baltic Sea, shifts in dominance from diatoms to dinoflagellates were linked to long-term warming trends (Kraberg et al., 2011). It follows that differential effects of warming can alter phytoplankton community composition and hence the roles of phytoplankton in nutrient and carbon cycling, food web dynamics and water quality (Scheffer et al., 2001; Elliott et al., 2005; Feuchtmayr et al., 2009; Jeppesen et al., 2010; Moss et al., 2011; Paerl and Paul 2011).

While these effects have focused on freshwater systems, estuarine and coastal systems are also affected (Paerl and Paul 2011; Kraberg et al., 2011), since freshwater and marine phytoplankton taxa strongly overlap in these systems. In addition, surface warming enhances vertical density stratification, especially in oligohaline waters. Changes in the strength, distribution and duration of stratification affect phytoplankton community structure, by favoring motile taxa such as dinoflagellates, other flagellated species and buoyant cyanobacteria over passive sinking taxa like diatoms (Reynolds 2006; Hall and Paerl 2011). Interestingly, phytoplankton groups containing harmful (i.e. toxic, food web disrupting) species, namely cyanobacteria and dinoflagellates, are favored selectively by warming effects, and warming appears responsible for the geographic expansion of toxic cyanobacterial bloom genera, including Anabaena, Cylindrospermopsis, Microcystis and Lyngbya. Examples include lakes in Northern Europe (Padisak 1997; Stuken et al. 2006; Wiedner et al., 2007) and the Baltic Sea (Suikkanen et al., 2007). Temperature regimes and relative cyanobacterial dominance were related positively for 146 lakes along a latitudinal gradient ranging from the sub-Arctic to southern South America (Kosten et al., 2012).

In estuarine and coastal benthic environments (seagrass beds, reefs, subtidal shelf and intertidal mudflats), filamentous attached cyanobacteria (Lyngbya spp., Oscillatoria spp.) are proliferating in systems that are impacted simultaneously by warming and nutrient enrichment (Paul 2008; Paerl and Paul 2011). Examples include Moreton Bay, Queensland, Australia (Watkinson et al., 2005; Ahern et al., 2007), coastal Florida (Paerl et al., 2008) and Guam (Kuffner et al., 2001). Detrimental effects include smothering of seagrass and coral communities, hypoxia, an increase in coral diseases (e.g., “black band disease”, caused by cyanobacteria), and declining finfish and shellfish habitats.

Climatic changes also affect magnitudes, geographic distributions and temporal patterns of precipitation (Trenberth 2005; Webster et al., 2005; O’Goman 2012. In some regions, both the amounts and extremes of precipitation have been altered. We are now experiencing record droughts and floods and changes in the frequency and intensity of tropical cyclones (Webster et al., 2005; Emanuel et al., 2008). These hydrologic changes impact phytoplankton production and bloom dynamics by: 1) altering, and as a result of extreme precipitation events, enhancing, nutrient loading by increasing erosion potentials and mobilizing land-based nutrients, 2) in the case of protracted droughts, increasing water residence time, which helps promote algal blooms, especially among slow-growing species (e.g., cyanobacteria, some dinoflagellates), 3) increasing water column stratification, which benefits motile/buoyant bloom-forming species (e.g., dinoflagellates, cyanobacteria), and 4) influencing the location and magnitude of phytoplankton production.

Hydrologic variability influences both the amounts and proportions of nutrients delivered to estuarine and coastal waters. The interactive effects of nutrient loading and hydrologic variability on estuarine phytoplankton production and community dynamics are particularly evident in poorly flushed (i.e., long water residence time) estuaries, which provide the opportunity to observe the fate of freshwater and nutrient inputs over spatial and time scales that overlap with those required for growth and bloom responses (i.e., relatively free of tidal flushing). Two neighboring semi-lagoonal estuaries located in Eastern North Carolina, the New and Neuse River Estuaries (Fig. 3), which are routinely monitored for water quality parameters, phytoplankton biomass and composition (Paerl et al., 2010; Peierls et al., 2012, Hall et al., 2013) provide an opportunity to examine these interactions.

Both systems have a history of anthropogenic (urban, agricultural and industrial) nutrient loading (Paerl et al., 1995; 2007; Mallin et al., 2005). They have also been under the influence of increasing hydrologic variability, including a recent increase in tropical cyclone activity interspersed with record droughts (Paerl et al., 2001, 2006; 2010; Peierls et al., 2012). Of the two systems, the Neuse River Estuary has had the longest intensive water quality and phytoplankton dynamics monitoring program in place (since the early 1990s) and hence is well suited for examining these interactions over a relatively long time frame (18 years). Here, high rainfall tropical cyclones and extratropical nor’easters have played a key role in modulating N and P loads to the estuary, as seen in Figure 4. Not surprisingly, storms that deposited high amounts of rainfall in the estuarine watersheds were associated with relatively high nutrient loads, while years when such events were absent (e.g., 1994) showed far lower seasonal and annual nutrient loads.

While total external N and P loads to both estuaries are controlled by freshwater discharge, the ratios of these nutrients are influenced by the magnitude of discharge. In both the Neuse and New River Estuaries, there is a strong positive relationship between DIN and total N concentrations and discharge under low to moderate discharge conditions (Fig. 5). Under elevated discharge conditions however, DIN and TN concentrations drop dramatically (Fig. 5), presumably due to dilution (Borsuk et al. 2004). In contrast, DIP and TP concentrations appear less strongly influenced by discharge over the entire range of discharge conditions (Fig. 5). In fact, the data suggest that at very high discharge conditions TP concentrations actually increase (Fig. 5). This suggests that the mobilization and transport dynamics of N and P fractions by freshwater discharge differ substantially under different flow regimes. Under relatively low discharge conditions, soluble forms of N and P tend to be most dominant, while under relatively high discharge conditions particulate fractions, including suspended sediments, play a larger role (data not shown). Under very high discharge conditions, total P concentrations tended to increase presumably due to the elevated suspended sediment load (from upstream soil and riverbed erosion) under these conditions. The result is that molar ratios of DIN to DIP and TN to TP (Fig. 5), are relatively low under low discharge, increase to maximum values during moderate discharge, and then fall under high discharge conditions.

These shifts in N:P supply ratios likely affect relative nutrient availabilities, limitation, and microalgal utilization/growth dynamics, which are known to influence inter-taxa competition and community structure in downstream waters (Tomas et al., 2007; Altman and Paerl 2012). This is a potentially-important, yet poorly understood aspect of how climatic changes may impacting estuarine and coastal eutrophication and phytoplankton/benthic microalgal community structure (including HABs).

Even though elevated nutrient loads associated with rainfall events provide resources for resident microalgal communities, the high freshwater discharge associated with these events also can advect phytoplankton cells downstream or even out of the estuary. Hence, the relative magnitudes of rainfall in these storms as well as their trajectory across the watershed and estuary were important determinants of the magnitudes as well as locations of phytoplankton biomass and bloom responses. Examples of several storm events are shown for the Neuse River Estuary, NC (Fig. 6).

In a most extreme hydrologic scenario, rainfall and flooding from sequential Hurricanes Dennis (10 d prior to Floyd), Floyd, and Irene (30 d after Floyd) impacted the Neuse River watershed over a period from late August through October 1999. The Neuse R. Estuary was flushed completely for more than two weeks following Floyd, leading to very low levels of resident phytoplankton throughout the estuary (Fig. 6). After about three weeks, phytoplankton biomass began to build in the lower estuary but increased flows due to Irene prevented bloom development. Throughout the two month long period following Dennis and Floyd, high flushing losses did not permit biomass accumulation in the upper half of the estuary. Phytoplankton growth rates were not able to catch up with flushing losses until late November when a phytoplankton bloom developed in the mid-estuary region (Fig. 6). Decreasing light availability from November through January prevented significant additional blooms from starting up in the estuary.

Hydrologic impacts of Hurricane Isabel (mid-September, 2003) were less severe. While Isabel was a powerful storm (Cat. 2), it contributed much smaller amounts of rainfall to the Neuse River watershed (< 20 cm) than the massive deluge (~ 1 m) that resulted from Floyd (1999) (Paerl et al., 2001). Prior to Isabel, phytoplankton blooms were present at upstream and mid-estuary stations (Fig. 6). Passage of this storm had little effect on phytoplankton biomass except a slight downstream shift in peak biomass (Fig. 6). Lack of a significant growth stimulation of biomass may have been due to the relatively high nutrient concentrations existing prior to the storm (Wetz and Paerl 2008).

Heavy rainfall from tropical storm Ernesto produced strong freshwater discharge to the Neuse R. Estuary in late August 2006 (Figs. 6 and 7A). Phytoplankton populations were largely flushed out of the upper reaches of the estuary as evidenced by the ~50% reduction of phytoplankton biomass at station 60 (Fig. 7 D). The freshwater inputs from the storm led to rapid increases in DIN in both surface and bottom waters of the upper estuary. A month-long period of intense, salinity-based vertical stratification (Fig. 7 B) allowed the accumulation of high levels of ammonium in the bottom waters (Fig. 7 C). As the storm waters receded and residence time increased, the combination of high nutrient availability and favorable residence times (Fig. 7 D) permitted bloom development by the toxic dinoflagellate, Karlodinium venificum. Cell concentrations reached 219,000 cells mL^-1 and the highly toxic cells were implicated in several fish kills in the area as the bloom crashed (Hall et al. 2008).

A further contrast was provided by Hurricane Irene (late August, 2011); whose eye also passed directly over the Pamlico Sound. The rain bands from this massive storm moved sufficiently inland to deliver large amounts of rainfall to the watershed. Prior to Irene, maximum phytoplankton biomass occurred in the upper regions (10-30 km downstream). After Irene, a well-defined peak in freshwater discharge elevated flushing (Fig. 6), which pushed the phytoplankton biomass peak downstream, but not out of the estuary as observed after Floyd. Once freshwater discharge subsided and residence times increased, phytoplankton biomass peaks resumed further upstream, where nutrient inputs were high.

Additional examples of how anthropogenic nutrient-driven eutrophication and algal bloom dynamics were modulated by climatically driven hydrologic variability can be shown for the New River Estuary (Hall et al., 2013; Peierls et al., 2012) (Fig. 8). This N-sensitive, microtidal system has exhibited a history of nutrient-enhanced primary production (Ensign et al., 2004; Mallin et al., 2005), harmful algal blooms (HABs) (Tomas et al., 2007), and bottom water hypoxia (Mallin et al. 2005). Phytoplankton biomass and composition are modulated strongly by variations in flow of the New River, the main freshwater input to the estuary, due to its influence on nutrient delivery and flushing time of the NRE. Phytoplankton biomass increased rapidly up to a threshold flushing time of ~10 d and then declined slowly at longer flushing times (Fig. 8)

This unimodal relationship indicated a balance between advective losses due to flushing and nutrient stimulation of biomass by riverine loading. Significant differences among group-specific optimal flushing times and rates of decline at longer flushing times suggest that hydrologic forcing plays an important role in determining phytoplankton composition and size structure (Hall et al. 2013; Paerl et al. 2013). Temperature additionally controlled phytoplankton composition. In particular, cyanobacteria abundance as the diagnostic photopigment zeaxanthin, which was dominated by picoplanktonic forms, showed a strong, positive relationship with water temperature and potentially toxic, bloom-forming raphidophytes occurred seasonally as temperatures warmed in late spring (Fig. 9).

Despite efforts to ameliorate eutrophication by reducing point source nutrient inputs through sewage treatment upgrades (Mallin et al., 2005), the New River is still impacted regularly by algal blooms, especially in the microtidal, upper estuarine region. Blooms are generally linked to elevated nutrient inputs in response to high flow periods. However, some blooms occur during droughts. This trend suggests that internal nutrient loading from the sediments may also play a critical role in bloom development. A positive feedback of phytoplankton biomass and sediment nutrient flux exists whereby blooms decrease light availability to the microphytobenthic community, decreasing benthic N demand, and increasing sediment N fluxes to the water column (Anderson et al. 2013).

Bloom forming flagellate species are particularly sensitive to changes in riverine nutrient inputs (Tomas et al., 2007; Altman and Paerl 2012). Sensitivity of the phytoplankton community was documented clearly when sewage treatment upgrades reduced nutrient loading to the New River Estuary by ~ 200,000 kg N yr^-1 and PP biomass fell by ~70 % (Mallin et al. 2005). Prior to sewage treatment upgrades, silica was, at times, potentially limiting (~ 0.5 µmol L^-1) for the growth of diatoms (Mallin et al. 1997) possibly explaining flagellate dominance of the NRE. Current silica concentrations (3 – 92 µmol L^-1) are unlikely to limit diatom growth (Dortch and Whitledge 1992), yet blooms are still dominated by flagellates, including some HAB species. It is likely that selective advantages gained by motility during strongly stratified periods (which are largely due to periods of elevated freshwater discharge), more than nutrient stoichiometry, explain why blooms in this estuary are dominated by flagellates. While nutrient reduction strategies may help reduce the magnitude of these blooms, density-driven stratification, which is largely attributable to precipitation and freshwater runoff conditions should also be considered.

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