THE H.T. ODUM SYNTHESIS ESSAY

Evolving Paradigms and Challenges in Estuarine and CoastalEutrophication Dynamics in a Culturallyand Climatically Stressed World

Hans W. Paerl & Nathan S. Hall & Benjamin L. Peierls &

Karen L. Rossignol

Received: 7 January 2014 /Accepted: 9 January 2014 /Published online: 6 February 2014# Coastal and Estuarine Research Federation 2014

Abstract Coastal watersheds support more than one half of theworld’s human population and are experiencing unprecedentedurban, agricultural, and industrial expansion. The freshwater–marine continua draining these watersheds are impacted increas-ingly by nutrient inputs and resultant eutrophication, includingsymptomatic harmful algal blooms, hypoxia, finfish and shell-fish kills, and loss of higher plant and animal habitat. In address-ing nutrient input reductions to stem and reverse eutrophication,phosphorus (P) has received priority traditionally in upstreamfreshwater regions, while controlling nitrogen (N) inputs hasbeen the focus of management strategies in estuarine and coastalwaters. However, freshwater, brackish, and full-salinity compo-nents of this continuum are connected structurally and function-ally. Intensification of human activities has caused imbalances inN and P loading, altering nutrient limitation characteristics andcomplicating successful eutrophication control along the contin-uum. Several recent examples indicate the need for dual N and Pinput constraints as the only nutrient management option effec-tive for long-term eutrophication control. Climatic changes in-crease variability in freshwater discharge with more severestorms and intense droughts and interact closely with nutrientinputs to modulate the magnitude and relative proportions of Nand P loading. The effects of these interactions on phytoplank-ton production and composition were examined in two neigh-boring North Carolina lagoonal estuaries, the New River andNeuse River Estuaries, which are experiencing concurrent eu-trophication and climatically driven hydrologic variability. Ef-forts aimed at stemming estuarine and coastal eutrophication inthese and other similarly impacted estuarine systems shouldfocus on establishing N and P input thresholds that take into

account effects of hydrologic variability, so that eutrophicationand harmful algal blooms can be controlled over a range ofcurrent and predicted climate change scenarios.

Keywords Nitrogen . Phosphorus . Hydrodynamics .

Phytoplankton . Coastal eutrophication . Nutrient limitation .

Climate change

Introduction

More than one half of the Earth’s human population resides incoastal water- and airsheds (Vitousek et al. 1997; NOAA2012; Ache et al. 2013). Accelerating agricultural, urban,and industrial development in these “sheds” has put unprece-dented pressure on the ecological condition and sustainabilityof downstream riverine, estuarine, and coastal waters (Brickeret al. 1999; Boesch et al. 2001; Conley et al. 2009). Nutrientover-enrichment has been identified as a prime cause for waterquality and habitat degradation (Nixon 1995; Paerl 1997;Boesch et al. 2001; Elmgren and Larsson 2001; Rabalaisand Turner 2001; Diaz and Rosenberg 2008). In addition,hydrologic modifications including upstream reservoir con-struction, and agricultural and urban surface and groundwaterwater withdrawal, have altered water flow rates and paths,sedimentation rates, and optical properties of receiving waters.These modifications affect estuarine and coastal water andhabitat quality (Cloern 2001; Boesch et al. 2001; Rabalaisand Turner 2001; Humborg et al. 2007). While human mod-ification of coastal water and airsheds directly, and oftennegatively, impacts water and habitat quality of these systems(Bricker et al. 1999; National Research Council 2000), cli-matic factors such as warming, more extreme storms, floods,and droughts modulate these impacts (Paerl and Huisman2008, 2009; Jeppesen et al. 2010).

Communicated by Iris C. Anderson

H. W. Paerl (*) :N. S. Hall : B. L. Peierls :K. L. RossignolUNC–CH Institute of Marine Sciences, Morehead City,NC 28557, USAe-mail: hans_paerl@unc.edu

Estuaries and Coasts (2014) 37:243–258DOI 10.1007/s12237-014-9773-x

Human and climatically induced stresses interact strongly inestuarine watersheds. These interactions affect the delivery andeffects of freshwater discharge and its nutrient load on theactivity and composition of microalgal primary producers ingeographically diverse estuarine and coastal ecosystems. Here,we evaluate the effects of human alterations of watershednutrients and hydrology on coastal eutrophication and the mea-sures needed to control it and associated habitat degradation in amore crowded and climatically extreme world. This issue isaddressed in the context of managing the freshwater to coastalcontinuum, and the physiographic and biogeochemical gradientcoupling water- and airsheds to coastal ecosystems (Fig. 1).

Nutrient Enrichment and Limitation in Estuarineand Coastal Ecosystems: Historical and CurrentPerspectives

The dominant nutrient limitation (of primary production) par-adigms applied to this continuum for more than half a centurywere that phosphorus (P) availability controls primary pro-duction in freshwaters, while nitrogen (N) is the dominantnutrient limiting production in the more saline downstreamestuarine and coastal waters (Ryther and Dunstan 1971;Schindler 1975; Nixon 1995; Boesch et al. 2001; Smith andSchindler 2009). Brackish waters often exhibit sensitivity toboth N and P inputs (Fisher et al. 1999; Rudek et al. 1991;Elmgren and Larsson 2001; Paerl and Piehler 2008). Recentanalyses of diverse nutrient limitation studies in both fresh-water and marine ecosystems indicate that these paradigmsmay be “eroding” (Elser et al. 2007; Lewis and Wurtsbaugh2008; Sterner 2008; Conley et al. 2009; Lewis et al. 2011).Increasingly, incidences of N and P “co-limitation”, i.e., thestimulation of primary production by the addition of N and P

in combination, where either N or P alone stimulate produc-tion far less, are observed (Elser et al. 2007; Lewis et al. 2011).Also, exclusive N limitation in freshwater ecosystems (asopposed to exclusive P limitation) is more common thanbelieved previously (Elser et al. 2007; Lewis et al. 2011).Concurrently, recent estuarine and coastal studies indicate thatN and P and/or P limitation are widespread geographically(Peeters and Peperzak 1990; Elmgren and Larsson 2001;Sylvan et al. 2006; Paerl and Justić 2011; Laurent et al. 2012).

Liebig’s “Law of the Minimum”: Theory Versus Practice

In examining nutrient limitation of primary producers,Liebig’s Law of the Minimum, also referred to as Liebig’sLaw or the Law of the Minimum, is an operational principlefirst 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 amountof resources available but by the scarcest resource (limitingfactor). From an aquatic plant production perspective, theyield is proportional to the amount of the most limiting re-source (i.e., nutrient, light, etc.). If the limiting resource is anutrient, it follows that yields may be improved by supplyingthe limiting nutrient to the point that some other nutrient isneeded in greater quantity than the water body can provide. Inthe case of N and P co-limitation, the balance between N and Psupply may approach the demand so that the addition of bothnutrients stimulates primary production, but separate additionsare not effective.

The application of Liebig’s Law was demonstrated inchemostat cultures with single phytoplankton species, wherenutrient supplies could be controlled tightly and growth yieldregulated by an essential nutrient supplied in amounts below

Fig. 1 Conceptual diagram, illustrating the complex interactions of human nutrient inputs, nutrient cycling processes, nutrient limitation (of primaryproduction) characteristics, and hydrologic forcing along the freshwater to marine continuum representing estuarine and coastal ecosystems

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those needed to maintain optimal cellular growth. Nutrientssupplied to these cultures can be manipulated at ratios andrates 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 shiftthe control of yield to the other nutrient. In natural systemssupporting complex phytoplankton communities, such nutri-ent supply shifts can occur as a result of variability in externalnutrient inputs (loads), internal nutrient cycling, and sedi-ment–water column nutrient exchange. Furthermore, changesin plant community composition, due to death, grazing, andplant–microbe and higher trophic level interactions (e.g., nu-trient regeneration from zooplankton grazers up to fish), alsoaffect nutrient availability and hence nutrient limitation. Thus,even subtle shifts in nutrient supply, community composition,and biogeochemical cycling can affect the nature and com-plexity of nutrient limitation.

Nutrient Limitation Paradigms

Historically, a much longer and geographically diverse line ofnutrient limitation data were available for freshwater thanmarine ecosystems, likely because the symptoms of nutrientover-enrichment and eutrophication were more evident andproblematic in freshwater ecosystems, dating back severalcenturies to the establishment and expansion of agriculture(i.e., rapid increase in chemical fertilizer use), the industrialrevolution, and urbanization (Thienemann 1915; Naumann1921; Parma 1980).

Freshwater and estuarine ecosystems often have largerwater- and airshed areas relative to their surfaces than do moreopen coastal marine ecosystems or large, deep lakes. Waterreplacement rate, relative to the volume of that water body(i.e., flushing rate), is often slow in shallow enclosed systems.Therefore, from a nutrient input and enrichment perspective,these systems are influenced heavily by their water- andairsheds. They tend to be dominated by N inputs since Ncompounds are often soluble and associated with a widevariety of organic and inorganic sources (e.g., plants, soils,atmospheric emissions and combustion products, and micro-bial transformations), whereas P is associated with rocks andsoils, where it is often insoluble and therefore unavailable.Furthermore, N, unlike P, exists in several oxidation states,including significant gaseous forms, and is more mobile andeasily transported and transformed in the geospheres, bio-spheres, and atmospheres. As such, freshwater to oligohalineestuarine systems are often enriched in N relative to P andexhibit P limitation.

Exceptions to this paradigm often relate to specific water-shed geochemical characteristics. Silicon (Si) may be deficientin watersheds that are dominated by non-silicious rocks andsoils (e.g., carbonates). This pattern can lead to Si limitation,

especially for diatoms, in downstream N- and P-enrichedwaters (Justić et al. 1995; Dortch and Whitledge 1992). Fur-thermore, construction of upstream dams and reservoirs canpromote “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 un-disturbed watersheds (cf. Wetzel 2001), supporting the earlysupposition that P is the limiting nutrient in most freshwaterecosystems (cf., Likens 1972). Unfortunately, only a few“undisturbed” air- and watersheds remain to evaluate thisparadigm. Today, agricultural, urban, and industrial expan-sions have altered the landscape, and amounts and patternsof nutrient loading to freshwater ecosystems. These activitieshave increased bothN and P loading, with wastewater inputsand runoff from land clearing and the establishment of farm-land and urban centers as dominant nutrient sources. The earlyrecognition of P as a primary limiting nutrient in these systems(Likens 1972; Schindler 1975), and the linkage of P loading tofreshwater eutrophication (Vollenweider 1968), provided theimpetus for focusing on P input reductions (Schindler andVallentyne 2008). Indeed, such reductions were effective instemming and reversing problematic symptoms of eutrophi-cation, nuisance algal blooms, food web disruption, bottomwater hypoxia, and degradation of planktonic and benthichabitats (Wetzel 2001; Schindler and Vallentyne 2008).

Nutrient loading dynamics have changed dramatically overthe past several decades. While P reductions were pursued,human population growth and agricultural and urban expan-sion in watersheds were paralleled by increased rates of Nloading (Peierls et al. 1991; Howarth 1998), often exceedingthose for P (Rabalais 2002; Galloway and Cowling 2002). Inthe Baltic Sea region, subjected to several centuries of humannutrient enrichment, effective control of eutrophication re-quires considering the total amounts and ratios of N and Pdischarged 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 banand improved wastewater treatment for P during the 1980s inNorth Carolina’s (USA) Neuse River System, helped arrestfreshwater blooms, but failure to reduce N inputs simulta-neously exacerbated blooms in downstream N-sensitive estu-arine waters (Paerl et al. 2004). In both cases, parallel N and Pinput reductions were required to stem eutrophication(Elmgren and Larsson 2001; Paerl 2009). In Florida’s (USA)extensive lake–river–estuary systems, excessive N loading,mainly from expanding wastewater and agricultural dis-charges, was identified (in addition to P) as supporting eutro-phication (Kratzer and Brezonik 1981). N2-fixingcyanobacteria often dominate in Lake Okeechobee, Florida’slargest lake. However, non-N2-fixing genera (e.g.,Microcystis) and facultative N2-fixing genera (e.g.,Cylindrospermopsis sp., Lyngbya sp.) compete effectively for

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reactive N when it is available and only fix N2 when otheravailable N is depleted (Moisander et al. 2012). In allcases, both N and P reductions are needed to controleutrophication 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 phyto-plankton dominated by non-N2-fixing cyanobacteria(Microcystis sp.; Lyngbya sp.) (cf., Steffen et al. 2014).Oligo- to mesohaline regions of large estuaries and coastalbays and seas (e.g., Chesapeake Bay, Albemarle–PamlicoSound, NC, Florida Bay, FL, Coastal North Sea and BalticSea;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). This co-limitation occurs because previ-ously loaded P and N are retained and recycled. This obser-vation is especially true for P, which exists only in a limitednumber of soluble (orthophosphate, dissolved organic P) andparticulate forms, which are cycled between the water columnand bottom sediments (Vollenweider 1968; Boynton andKemp 1985; Wetzel 2001).

In contrast, N exists in multiple forms, including dissolved(nitrate, nitrite, ammonium, organic N), particulate, and gas-eous; P has no significant gaseous forms. Gaseous forms of N(N2, N2O, NO, NO2, NH3, volatile organic N compounds) areproduced by microbial transformations, including ammonifi-cation, anammox, nitrification, and denitrification (Caponeet al. 2008), which can escape into the atmosphere. In partic-ular, 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 fuelcombustion, wastewater treatment, and a wide range of indus-trial chemical processes are major sources of biologicallyreactive N to estuarine and coastal waters, and have increaseddramatically over the past half-century (Galloway andCowling 2002; US EPA 2011). In Northern Gulf of Mexicowaters receiving discharge from the Mississippi River Basin,agricultural and urban N inputs in the Basin have increased sorapidly that the receiving marine waters now exhibit P limita-tion during spring with elevated runoff, while N limitationprevails during the drier summer months (Sylvan et al. 2006).Anthropogenically generated P has also increased, but inmany instances not nearly as rapidly as N (Justić et al.1995), especially in places where P detergent bans and im-proved wastewater treatment for P have been implemented. Inmany intensively farmed, urbanized, and industrialized re-gions, however, historic and current P loads are still quite highbecause of P-saturated soils and continued P fertilizer

applications. As a result, residual P supplies in sediments haveremained high and available.

While a N input “glut” is occurring due to expandinganthropogenic inputs, a fraction of the N supplied to waterbodies is “lost” as N2 via denitrification (Seitzinger 1988) orconverted to other gaseous forms (e.g., NH3, N2O, and NOemissions) (US EPA 2011). N2 fixation rates are not sufficientto offset N losses via denitrification and anammox in thesesystems, perpetuating N-limited conditions (Paerl and Scott2010). Thus, despite receiving ever-increasing anthropogenicN inputs, these systems can still assimilate such inputs andbecome more eutrophic, without becoming exclusively Plimited, due to N losses via denitrification and other gas-generating processes, while much of the P is internally cycled.This phenomenon appears widespread in meso- to eutrophicfreshwater and marine ecosystems, which exhibit N limitationor N and P co-limitation (Granéli et al. 1999; Elmgren andLarsson 2001; Elser et al. 2007; Sterner 2008; Paerl andPiehler 2008; Finlay et al. 2010; Lewis et al. 2011).

Interestingly, nutrient management efforts in freshwatercomponents of the estuarine continua continue to focus largelyon “P only” reduction strategies (cf., Schindler and Vallentyne2008; Schindler et al. 2008), despite pioneering studies in the1960s showing a role for N in freshwater eutrophication (cf.,Goldman 1981; Wetzel 2001), and more recent studies dem-onstrating sensitivity of a range of lakes and reservoirs to Ninputs (Elser et al. 2007; North et al. 2007; Lewis andWurtsbaugh 2008; Finlay et al. 2010; Xu et al. 2010; Paerlet al. 2011b; Spivak et al. 2011; Lewis et al. 2011). Thisapproach is based on the assumption that cyanobacteria cansupply N via nitrogen (N2) fixation (Schindler et al. 2008).However, at the ecosystem level, only a fraction, usually farless 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 Ndrawdown (Scott et al. 2009). This chronic N deficit appearsto be even more pronounced for estuarine and coastal waters(Howarth et al. 1988). Indications are that N2 fixation iscontrolled by factors in addition to just P availability (Paerl1990; Scott and McCarthy 2010). N-limitation may persist inaquatic ecosystems, even in the presence of N2 fixers. There-fore, external N inputs play a key role in controlling primaryproduction along the continuum (cf., McCarthy et al. 2007;Conley et al. 2009; Paerl 2009; Paerl et al. 2011b).

From a nutrient management perspective, important ques-tions and research needs include (1) How do patterns ofreactive N drawdown lead to seasonal N limitation or N + Pco-limitation? (2) How is this drawdown partitioned betweenphytoplankton N demand and denitrification? (3) What is thecritical balance (i.e., threshold) between N/P and loading ratesand removal processes (i.e., uptake, denitrification) over epi-sodic, seasonal, and multi-annual time scales?

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The Interacting Roles of Climate Change: Warmingand Increased Hydrologic Variability

Climatic changes, specifically warming and altered rainfallamounts and patterns, interact with nutrient enrichment inmodulating eutrophication dynamics. Temperature controlsmicrobial metabolism and hence carbon and nutrient cycling.In addition, it controls microalgal growth rates, and the con-trols appear taxa-specific (cf., Paerl et al. 2011a) (Fig. 2).Mostnotable is the stimulatory effect of warming on growth rates ofcyanobacteria, the only prokaryotic phytoplankton group(Paerl and Huisman 2008). Being “bacterial”, cyanobacterialspecies tend to show growth optima that are in the 25–30 °Crange, in contrast to eukaryotic taxa, which typically exhibitgrowth maxima at lower temperatures (Paerl et al. 2011a).Therefore, longer, warmer growing seasons favorcyanobacterial species (Jōhnk et al. 2008; Paerl et al. 2011a).Also, there are taxa-specific effects of warming on eukaryoticphytoplankton groups. For example in the Baltic Sea, shifts indominance from diatoms to dinoflagellates were linked tolong-term warming trends (Kraberg et al. 2011). It followsthat differential effects of warming can alter phytoplanktoncommunity composition and hence the roles of phytoplanktonin nutrient and carbon cycling, food web dynamics, and waterquality (Scheffer et al. 2001; Elliott et al. 2005; Feuchtmayret al. 2009; Jeppesen et al. 2010; Moss et al. 2011; Paerl andPaul 2011).

While these studies have largely focused on freshwatersystems, estuarine and coastal systems are also affected(Paerl and Paul 2011; Kraberg et al. 2011) since freshwaterand marine phytoplankton taxa overlap in these systems. Inaddition, surface warming enhances vertical density stratifica-tion, especially in oligohaline waters. Changes in the strength,

distribution, and duration of stratification affect phytoplank-ton community structure by favoring motile taxa such asdinoflagellates, other flagellated species, and buoyantcyanobacteria over passive sinking taxa like diatoms(Reynolds 2006; Hall and Paerl 2011). Interestingly, phyto-plankton groups containing harmful (i.e., toxic, food webdisrupting) species, namely cyanobacteria and dinoflagellates,are favored selectively by warming effects, and warmingappears responsible for the geographic expansion of toxiccyanobacterial bloom genera, including Anabaena,Cylindrospermopsis, Microcystis, and Lyngbya. Examples in-clude lakes in Northern Europe (Padisak 1997; Stüken et al.2006;Wiedner et al. 2007) and the Baltic Sea (Suikkanen et al.2007). Temperature regimes and relative cyanobacterial dom-inance were related positively for 146 lakes along a latitudinalgradient ranging from the sub-Arctic to southern South Amer-ica (Kosten et al. 2012).

In estuarine and coastal benthic environments (seagrassbeds, reefs, subtidal shelf, and intertidal mudflats), filamen-tous attached cyanobacteria (Lyngbya spp., Oscillatoria spp.)are proliferating in systems that are impacted simultaneouslyby warming and nutrient enrichment (Paul 2008; Paerl andPaul 2011). Examples include Moreton Bay, Queensland,Australia (Watkinson et al. 2005; Ahern et al. 2007), coastalFlorida (Paerl et al. 2008), and Guam (Kuffner and Paul2001). Detrimental effects include smothering of seagrassand coral communities, hypoxia, an increase in coral diseases(e.g., “black band disease”, caused by cyanobacteria), anddeclining finfish and shellfish habitats.

Climatic changes also affect magnitudes, geographic dis-tributions, and temporal patterns of precipitation (Trenberth2005; Webster et al. 2005; O’Goman 2012. In some regions,both the amounts and extremes of precipitation are altered, as

0 5 10 15 20 25 30 35 40

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Fig. 2 Temperature dependenceof the specific growth rates ofrepresentative species from threeeukaryotic phytoplankton classesand of bloom-formingcyanobacterial species commonin temperate freshwater andbrackish environments. Datapoints represent a 5 °C movingaverage of from laboratorygrowth experiments under light-and nutrient-saturated conditions.References for the individualspecies can be found in Paerl et al.(2011b). Solid lines are a best-fittemperature optimum function(Chen and Millero 1986)

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evidenced by record droughts and floods and changes in thefrequency and intensity of tropical cyclones (Webster et al.2005; Emanuel et al. 2008). These hydrologic changes impactphytoplankton production and bloom dynamics by (1) alter-ing, and as a result of extreme precipitation events, enhancing,nutrient loading by increasing erosion potentials and mobiliz-ing land-based nutrients; (2) in the case of protracted droughts,increasing water residence time, which helps promote algalblooms, especially among slower-growing species (e.g.,cyanobacteria, some dinoflagellates); (3) increasing water col-umn stratification, which benefits motile/buoyant bloom-forming species (e.g., dinoflagellates, cyanobacteria); and (4)influencing the location and magnitude of phytoplanktonproduction.

Hydrologic variability influences both the amounts andproportions of nutrients delivered to estuarine and coastalwaters. The interactive effects of nutrient loading and hydro-logic variability on estuarine phytoplankton production andcommunity dynamics are particularly evident in poorlyflushed (i.e., long water residence time) estuaries, which pro-vide the opportunity to observe the fate of freshwater andnutrient inputs over spatial and time scales that overlap withthose required for growth and bloom responses (i.e., relativelyfree of tidal flushing). Two neighboring semi-lagoonal estu-aries located in Eastern North Carolina, the New and NeuseRiver Estuaries (Fig. 3), provide an opportunity to examinethese interactions. Both rivers are monitored routinely forwater quality parameters, and phytoplankton biomass andcomposition (Paerl et al. 2010; Peierls et al. 2012; Hall et al.2013).

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 influ-enced by increasing hydrologic variability, including a recentincrease in tropical cyclone activity interspersed with record

droughts (Paerl et al. 2001, 2006, 2010; Peierls et al. 2012).The Neuse River Estuary has had an intensive water qualityand phytoplankton dynamics monitoring program in placesince the early 1990s and hence is well suited for examiningthese interactions over a relatively long time frame (18 years).High rainfall tropical cyclones and extratropical nor’eastershave played a key role in modulating N and P loads to theestuary (Fig. 4). Not surprisingly, storms that deposited highamounts of rainfall in the estuarine watersheds were associat-ed with high nutrient loads. Far lower seasonal and annualnutrient loads occurred during years without such events (e.g.,1994).

Total external N and P loads to both estuaries are controlledby freshwater discharge, and the ratios of these nutrients areinfluenced by the magnitude of discharge. A strong positiverelationship between dissolved inorganic nitrogen (DIN) andtotal N concentrations and discharge under low to moderatedischarge conditions occurs in both the Neuse and New RiverEstuaries (Fig. 5). However, DIN and TN concentrations dropdramatically under elevated discharge conditions (Fig. 5),presumably from dilution (Borsuk et al. 2004). DIP and TPconcentrations are also strongly influenced by discharge overthe entire range of discharge conditions (Fig. 5). In fact, TPconcentrations actually increase under very high dischargeconditions (Fig. 5). This suggests that the mobilization andtransport dynamics of N and P fractions by freshwater dis-charge differ substantially under different flow regimes. Sol-uble forms of N and P are most dominant under low dischargeconditions, while particulate fractions, including suspendedsediments, play a larger role under high discharge conditions(data not shown). Total P concentrations tended to increaseunder very high discharge conditions, presumably due to theelevated suspended sediment load (from upstream soil andriverbed erosion) under these conditions. As a result, molarratios of DIN to DIP and TN to TP are relatively small under

Fig. 3 Maps of the aNew River Estuary and bNeuse River Estuary, NC,showing the locations of sampling stations and US Geological Service(USGS) river gaging stations. The USGS gaging station on the Neuse

River is located at Fort Barnwell approximately 26 km upstream fromNew Bern and out of the area covered by (b)

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low discharge, increase to maximum values during moderatedischarge, and then fall under high discharge conditions(Fig. 5).

These shifts in N/P supply ratios likely affect relativenutrient availabilities, limitation, and microalgal utilization/growth dynamics, which can influence inter-taxa competitionand community structure in downstream waters (Tomas et al.2007; Altman and Paerl 2012). These interactions arepotentially important yet poorly understood examples of howclimatic changesmay impact estuarine and coastal eutrophication

and phytoplankton/benthic microalgal community structure (in-cluding harmful algal blooms, HABs).

The effects of rainfall events on coastal ecosystems arecomplex and interactive. For example, elevated nutrient loadsassociated with freshwater discharge provide resources forresident microalgal communities; however, high flow canadvect phytoplankton cells downstream or even out of theestuary. Hence, the relative magnitudes of rainfall in thesestorms as well as their trajectory across the watershed andestuary are important determinants of the magnitudes and

Fig. 4 Dissolved inorganicnitrogen (a) and dissolvedinorganic phosphorus (b) loading,as metric tons per day, enteringthe Neuse River Estuary duringyears having no major tropicalcyclones impacting its watershed(1994) versus years in whichstorms (named) impacted thewatershed. Note the variability innutrient loads associated withspecific storms, which waslargely attributed to the stormtracks relative to the watershedand the amount of rainfalldeposited by each storm

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locations of phytoplankton biomass and bloom responses.Examples of several storm events are shown for the NeuseRiver Estuary, NC (Fig. 6).

In a most extreme hydrologic scenario, rainfall andflooding from sequential Hurricanes Dennis (10 days priorto Floyd), Floyd, and Irene (30 days after Floyd) impacted theNeuse River watershed over a period from late Augustthrough October 1999. The Neuse River Estuary was flushedcompletely for more than 2 weeks following Floyd, leading tovery low levels of resident phytoplankton throughout theestuary (Fig. 6). After about 3 weeks, phytoplankton biomassbegan to build in the lower estuary, but increased flows due toIrene prevented bloom development. Throughout the 2-monthperiod following Dennis and Floyd, high flushing losses didnot permit biomass accumulation in the upper half of theestuary. Phytoplankton growth rates did not catch up withflushing losses until late November when a phytoplanktonbloom developed in the mid-estuary region (Fig. 6). Decreas-ing light availability from November to January preventeddevelopment of significant additional blooms in the estuary(data not shown).

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 theNeuse 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 up-stream and mid-estuary stations (Fig. 6). Passage of this stormhad little effect on phytoplankton biomass except for a slightdownstream shift in peak biomass (Fig. 6). Lack of a signif-icant growth stimulation of biomass may have resulted fromthe high nutrient concentrations existing prior to the storm(Wetz and Paerl 2008).

Heavy rainfall from tropical storm Ernesto produced strongfreshwater discharge to the Neuse River Estuary in late Au-gust 2006 (Figs. 6 and 7a). Phytoplankton populations werelargely flushed out of the upper reaches of the estuary asevidenced by the ∼50 % reduction of phytoplankton biomassat station 60 (Fig. 7d). The freshwater inputs from the stormled to rapid increases in DIN in both surface and bottomwaters of the upper estuary. A month-long period of intense,salinity-based vertical stratification (Fig. 7b) allowed highlevels of ammonium to accumulate in the bottom waters(Fig. 7c). As the storm waters receded and residence timeincreased, the combination of high nutrient availability andfavorable residence times (Fig. 7d) permitted bloom develop-ment by the toxic dinoflagellate, Karlodinium venificum. Cellconcentrations reached 219,000 cells mL−1 and the highlytoxic cells were implicated in several fish kills in the area asthe bloom crashed (Hall et al. 2008).

A further contrast was provided by Hurricane Irene (lateAugust, 2011), whose eye also passed directly over the Pam-lico Sound. The rain bands from this massive storm moved

sufficiently inland to deliver large amounts of rainfall to thewatershed. Prior to Irene, maximum phytoplankton biomassoccurred in the upper regions (10–30 km downstream). AfterIrene, a well-defined peak in freshwater discharge increasedflushing (Fig. 6), which pushed the phytoplankton biomasspeak downstream, but not out of the estuary as observed afterFloyd. Once freshwater discharge subsided and flushing ratesdecreased, phytoplankton biomass peaks resumed further up-stream, where nutrient inputs were high.

Additional examples of how anthropogenic nutrient-driveneutrophication and algal bloom dynamics were modulated byclimatically driven hydrologic variability were shown for theNew River Estuary (NRE) (Hall et al. 2013; Peierls et al.2012) (Fig. 8). This N-sensitive, microtidal system has exhib-ited a history of nutrient-enhanced primary production(Ensign et al. 2004; Mallin et al. 2005), HABs (Tomas et al.2007), and bottom water hypoxia (Mallin et al. 2005). Phyto-plankton biomass and composition are modulated by varia-tions in flow of the New River, the main freshwater input tothe estuary, due to its influence on nutrient delivery andflushing time of the NRE. Phytoplankton biomass increasedrapidly up to a threshold flushing time of ∼10 days and thendeclined slowly at longer flushing times (Fig. 8).

This unimodal relationship indicated a balance betweenadvective losses due to flushing and nutrient stimulation ofbiomass by riverine loading. Significant differences amonggroup-specific optimal flushing times and rates of declineduring longer flushing times suggest that hydrologic forcingis important to determining phytoplankton composition andsize structure (Hall et al. 2013; Paerl et al. 2013). Temperaturealso controlled phytoplankton composition. The diagnosticphotopigment zeaxanthin, primarily associated withpicoplanktonic forms such as cyanobacteria, showed a strong,positive relationship with water temperature. The potentiallytoxic, bloom-forming raphidophytes also occurred seasonallyas temperatures warmed in late spring (Fig. 9).

Despite efforts to ameliorate eutrophication by reducingpoint source nutrient inputs through sewage treatment up-grades (Mallin et al. 2005), the New River is still impactedregularly by algal blooms, especially in the microtidal, upperestuarine region. Blooms are generally linked to elevatednutrient inputs in response to high flow periods. However,some blooms occur during droughts. This trend suggests thatinternal nutrient loading from the sediments may also play acritical role in bloom development. A positive feedback ofphytoplankton biomass and sediment nutrient flux exists

Fig. 5 Concentrations and molar ratios of nitrogen and phosphorusversus river flow in the Neuse and New Rivers just above the limit ofsalt intrusion. River flow and nutrient concentrations were measured atthe USGS Gum Branch gaging station for the New River. Neuse Riverflow was measured at Fort Barnwell and nutrient concentrations weremeasured at Streets Ferry Bridge

�

250 Estuaries and Coasts (2014) 37:243–258

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whereby blooms decrease light availability to themicrophytobenthic community, decreasing benthic N de-mand, and increasing sediment N fluxes to the water column(Anderson et al. 2013).

Bloom-forming flagellate species are sensitive to chang-es in riverine nutrient inputs (Tomas et al. 2007; Altmanand Paerl 2012). Sensitivity of the phytoplankton commu-nity was documented when sewage treatment upgradesreduced nutrient loading to the New River Estuary by∼200,000 kg N year−1 and PP biomass fell by ∼70 %(Mallin et al. 2005). Prior to sewage treatment upgrades,silica, at times, potentially limited (∼0.5 μmol L−1) thegrowth of diatoms (Mallin et al. 1997), possibly explainingflagellate dominance of the NRE. Current silica concentra-tions (3–92 μmol L−1) are unlikely to limit diatom growth(Dortch and Whitledge 1992), yet blooms are still domi-nated by flagellates, including some HAB species.

While nutrient reduction strategies may help reduce themagnitude of these blooms, density-driven stratification,

which is largely attributable to precipitation and freshwaterrunoff conditions, also likely plays an important role in deter-mining phytoplankton community structure. Selective advan-tages gained by motility during stratified periods may explainwhy blooms in this estuary are dominated by flagellates.

Closing Remarks

Synergistic anthropogenic nutrient loading and climaticchanges, including warming, enhanced vertical stratification,and periods of increased flushing time, provide the mecha-nisms for accelerating eutrophication and promoting an in-crease in HAB frequency and intensity in estuarine and coastalwaters (Fig. 10). Nutrient input reductions are the most obvi-ous “knobs” that can be “tweaked” to reduce estuarine andcoastal eutrophication, and as such should be the key compo-nent of most management strategies. We have long been

Tropical Storm Ernesto 31 August 2006

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Fig. 6 Impacts of four tropical cyclones on river flow and the down-stream distribution of chlorophyll a in the Neuse River Estuary. Dates foreach storm represent landfall on the North Carolina coast. Letters abovethe flow panel correspondwith the letterson the chlorophyll adistributionpanels. Off-scale chlorophyll a values are written below the peaks.

Extremely high flow events, such as the one following Hurricane Floyd(1999), while delivering large amounts of nutrients, also led to “washout”of resident phytoplankton communities, thereby negating the potentialstimulatory effect of this event

252 Estuaries and Coasts (2014) 37:243–258

aware of the role P inputs play in such a strategy. However,increasing evidence suggests that N input reductions are alsoneeded. A dual N and P reduction strategy offers the mostrealistic and effective long-term approach to eutrophicationmanagement along the estuarine continuum. A key manage-ment priority is to establish N and P input thresholds [e.g., theUSEPA’s Total Maximum Daily Loads or TMDLs; the Euro-pean Union Water Framework Directive—(http://ec.europa.eu/environment/water/index_en.htm)], below whicheutrophication and HABs can be controlled in terms ofmagnitude, and temporal and spatial coverage. Thesethresholds should incorporate the effects of hydrologic and

temperature conditions, predicted to vary in response toclimate change, as these factors affect the ratio of N/P loadsand modulate estuarine and coastal responses to nutrientinputs.

Both the total amounts and ratios of N to P inputs should beconsidered when developing these thresholds. Ideal inputratios should not favor specific HAB taxa over others, butunfortunately a universal ratio—above or below—to controlHABs consistently and reliably is not available. Thus, system-specific input ratios are needed for appropriate and effectivecontrol. Total molar N/P ratios above ∼15 may discourageCyanoHAB dominance (cf., Smith and Schindler 2009).

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Fig. 7 Environmental conditions leading to the development of a toxicKarlodinium veneficumbloom in the Neuse River Estuary. aNeuse Riverflow at Fort Barnwell, NC showing the runoff pulse from Tropical StormErnesto during the fall of 2006. bTime series of surface (0 m) and bottom(3 m) salinity from automated USGS instrumentation at station CM11,∼4 km downstream of station 60 where peak bloom concentrations were

observed on 19 October. c Time series of surface and bottom dissolvedinorganic nitrogen at station 60. Inset pie graphs below time series linesshow the relative proportion of nitrate (black fill) versus ammonium(white fill) in the DIN pool of surface and bottom waters on 3 October.dTime series of surface water chlorophyll a,K. veneficumcell abundance,and flushing time at station 60 (see Fig. 3 for its location)

Estuaries and Coasts (2014) 37:243–258 253

However, if the nutrient concentrations in receiving waters ofN or P exceed uptake saturation values, a ratio approach forreducing eutrophication and HAB formation may not beeffective.

There are many ways to reduce nutrient inputs onecosystem-specific scales. Nutrient inputs are classified aspoint source and non-point source. Point sources are associ-ated with well-defined and identifiable discharge sites; there-fore, these nutrient inputs are often considered “low hangingfruit”, i.e., relatively easy to control. Most short-term suc-cesses in nutrient input control were accomplished with pointsource reductions of wastewater, industrial effluents, and otherdistinct input sources. The major remaining challenge in manywatersheds is targeting and controlling non-point sources,

which frequently are the largest sources of nutrientsdischarged to coastal waters (National Research Council2000; US EPA 2011); their controls are critical in mitigatingeutrophication and HABs.

Manipulating physical factors that play key roles in con-trolling the composition and function of phytoplankton (andbenthic microalgal) communities will benefit systems sensi-tive to eutrophication and HABs. Vertical mixing devices,bubblers, and other means of destratification are effective incontrolling outbreaks and persistence of some HABs (e.g.,cyanobacteria), but only in relatively small systems (cf.,Hudnell 2008). These devices have limitations in estuarineand coastal waters because they are not applicable to largeareas and volumes. Furthermore, artificial mixing does not

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Fig. 9 Seasonality of picocyanobacteria and raphidophyte blooms in theNew River Estuary. Contour plot of zeaxanthin concentration versus timeand distance upstream from the New River Inlet with average tempera-tures from all stations (solid black line, error bars are standard deviation).

White circles show the time and spatial locations of raphidophyte bloomsdefined as chlorophyll a concentrations greater than 40 μg L−1 andraphidophytes comprising more than half of total phytoplanktonbiovolume

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254 Estuaries and Coasts (2014) 37:243–258

mitigate the underlying problem of nutrient over-enrichment.In fact, this approach may be counterproductive in verticallystratified waters, by transporting bottomwater nutrients acrossthe pycnocline up to the surface, potentially exacerbatingsurface-dwelling blooms.

Increasing the flushing rates can reduce or control HABs inthese systems (cf., Paerl et al. 2011b). However, the flushingwater must have low nutrient content to prevent further en-richment. Algal community structuring effects of changingN/P ratios, which can take place as freshwater discharge isaltered, must be considered. Furthermore, few communitiescan afford to use precious water resources normally reservedfor drinking or irrigation water for flushing purposes, espe-cially in regions with limited or drought-impacted freshwaterrunoff. Lastly, flushing can alter the circulation regimes ofestuarine and coastal waters. Care must be taken to preventtrapping of the HABs in the system by altering the physicalenvironment (e.g., increasing thermal or chemical densitystratification, entrainment bays, and arms of water bodies),rather than flushing them out of the system.

Nutrient input reductions are in general the most direct,simple, ecologically rational, and economically feasible eutro-phication and HAB management strategy. Nutrient input re-ductions aimed specifically at reducing HAB competitive abil-ities, and possibly combined with physical controls (in systemsamenable to those controls), are often the most effective strat-egies. An obvious strategy which is gaining traction is apply-ing fertilizers at “agronomic rates”, i.e., satisfying crop needs,while avoiding excesses and modifying drainage ditches andtile drains to increase their efficiency in minimizing nutrientlosses (David et al. 2010). Nutrient (specifically N) removalfrom wastewater can be prohibitively expensive, so that alter-native nutrient removal strategies are needed. Alternate strate-gies may include construction of wetlands, cultivation andstimulation of macrophytes, and stocking of herbivorous (spe-cifically cyanobacteria consumers) fish and shellfish species.

In addition to nutrient input reductions, future water man-agement strategies will need to accommodate the hydrological

and physical–chemical effects of climatic change. Withoutcurbing greenhouse gas emissions, future warming trendsand hydrological extremeness will degrade estuarine andcoastal ecosystem water and habitat quality further, increasingthe potential for expansion by opportunistic nuisancemicroalgae and cyanobacteria.

Acknowledgments We thank co-workers who assisted with field andlaboratory work and manuscript preparation, including B. Abare,J, Braddy, A. Joyner, L. Kelly, and R. Sloup. The editorial inputfrom I. Anderson andW. Gardner and review provided byK.McGlatheryare appreciated. This research was funded by the Strategic EnvironmentalResearch and Developmental Program (SERDP)–Defense Coastal/Estuarine Research Program, Project SI-1413, The Lower Neuse BasinAssociation/Neuse River Compliance Association, the North CarolinaDepartment of Environment and Natural Resources (ModMon Program),and National Science Foundation Projects OCE 0825466, OCE 0812913,ENG/CBET0826819 and 1230543, and DEB 1119704 and 1240851.

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CHARACTERISTICS

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