Eutrophication in Little Pond: Algal Growth Response to Nutrients and Immediate Effects

of Sewer Systems on Enrichment

Briana Moore

University of Chicago

Chicago, IL 60637 USA

Abstract

Little Pond, an estuary in Falmouth, Massachusetts, has faced elevated nitrogen (N) levels for years, largely as a result of wastewater loading from nearby septic systems. The installation of a sewer system to redirect the wastewater to a treatment plant has the potential to reverse the eutrophication of the water in the pond. Still, although estuaries are usually N- limited, high N loading has the potential to shift primary limitation to another nutrient, phosphorus (P). I investigated this limitation by studying the following: (1) the amount of nutrients entering the pond through groundwater; (2) the concentration of nutrients (dissolved - + -3 inorganic nitrogen, DIN = [NO3 + NH4 ], and dissolved inorganic phosphorus, DIP = [PO4 ]) in the body of the estuary across a range of depths and salinities; and (3) how much the addition of nitrogen and/or phosphorus influences the growth of phytoplankton in water of varying salinities collected from the estuary. I found average DIN concentrations of 116 μmol/L in the groundwater and 152 μmol/L in the stream, and average DIP concentrations of 2.8 μmol/L in the groundwater and 1.7 μmol/L in the stream. I also found more intense stratification in the freshwater sites compared to mixing in the saltier sites within the estuary. Finally, I conducted a bioassay of surface water and measured chlorophyll and nutrient concentrations over a several- day incubation period, finding that DIP was limiting in lower-salinity waters (0–18 ‰) and that DIP and DIN were co-limiting in higher-salinity waters (27 ‰). Peak chlorophyll a (in-vivo fluorescence) and POC levels were 5.6 μg/L and 312 μmol/L respectively at 0 ‰; 14.8 μg/L and 599 μmol/L at 8 ‰; 19.8 μg/L and 508 μmol/L at 18 ‰; and 17.6 μg/L and 526 μmol/L at 27 ‰. Altogether, this investigation showed that Little Pond receives so much N loading that P has replaced N as the limiting nutrient.

Key Words and Phrases

Eutrophication, nitrogen and phosphorus contamination, algal nutrient limitation in estuaries.

Introduction

Eutrophication, the overgrowth of algae and plants usually caused by excessive amounts of nutrients supplied to a body of water, has become a more common problem in recent decades around the world (Nixon 2009, Rabalais et al. 2009). Population growth, manufacture and use of fertilizer for agriculture, and fossil fuel combustion have increased loading of nutrients, nitrogen in particular, to coastal bays and estuaries, encouraging primary producers to accumulate biomass at unsustainable rates and leading to an increased presence of hypoxic and anoxic zones, fish kills, loss of biodiversity, harmful algal blooms, and poor water quality (Diaz and Rosenberg 2008, Rabalais et al. 2009). Not only do worsened water conditions harm recreation and tourism in these areas; they can cause entire ecosystems to collapse (Diaz and Rosenberg 2008). Reducing nutrient loading to coastal aquatic systems is a more urgent endeavor as a result, and even more so as experts speculate that eutrophication could work hand-in-hand with warming global temperatures and more powerful storms to exacerbate these effects (Moss et al. 2011).

Little Pond, a coastal embayment in Falmouth, Massachusetts (Figure 1), has suffered from enriched loading for years; the Massachusetts Estuaries Project considered the estuary’s ecological condition “significantly impaired” (Cape Cod Commission 2017). The conversion of surrounding forest into residences served by on-site septic systems for a swelling local population has caused major increases in nitrogen loaded into the estuary through controllable means, leading to water quality impairment (Howes et al. 2006). In fact, septic systems and other controllable sources of nitrogen loading make up 86% of total loading to Little Pond (Howes et al. 2006).

In response, the Town of Falmouth has installed a sewer system designed to collect and treat wastewater from 1,440 homes in the Little Pond watershed (Figure 1, right panel). The system will redirect contaminated wastewater away from the pond to the Falmouth Wastewater Treatment Facility and is expected to reduce the N and P concentrations in the groundwater entering Little Pond.

A study by the Massachusetts Estuaries Project (Howes et al. 2006) recommended that the total maximum daily N loading (TMDL) to Little Pond to restore a healthy be no more than 7 kg/day, far below both septic systems loading and other controllable source loading, let alone total loading (Figure 2). The Massachusetts Department of Environmental Protection (DEP) has adopted this target as a regulatory threshold.

By measuring current contaminant concentrations in the groundwater entering the system and comparing them with similar data collected before the installation of the sewer system, I offer insight into the system’s effectiveness, at least with regard to changes in nutrient levels within the estuary. The legacy of past contaminant influx means it may take several years for substantial changes to manifest, so sampling at this point in the project provides a baseline to be compared with future observations. To document baseline pre-sewering groundwater nutrient levels, nitrogen and phosphorus have been measured in water collected from wells along the eastern and western shore of Little Pond five times in the past three years (personal communication, Cropper and Foreman 2015, Goodman and Foreman 2016; Earisman and Foreman 2016; Benning-Shorb 2017; personal communication, Duran and Foreman 2018). Benning-Shorb (2017) also measured nitrogen and phosphorus influxes through groundwater, outfluxes through the channel into Vineyard Sound, and sediment N and P mineralization and flux before the activation of the sewer system. Since he completed his work, over 1,000 homes have been connected to the sewer system (personal communication, Town of Falmouth wastewater superintendent Amy Lowell). In this study, I continue to make baseline measurements to estimate the nutrient concentrations in the estuary and predict how they may change after the installation of the sewer system. Prior work has shown that nitrogen inputs are elevated so that the N:P ratio in the groundwater and stream water entering Little Pond exceeds the Redfield Ratio, 16N:1P.

I also conduct bioassays to evaluate whether the high rates of nitrogen input have shifted the limiting nutrients towards phosphorus, despite estuaries like Little Pond being naturally nitrogen-limited (Howarth and Marino 2006). Still, it has been shown that excessive supplies of nitrogen can shift N:P ratios and change N-limitation to P-limitation (Elser et al. 2012). I collect surface water samples from the freshwater stream to the north and down the salinity gradient to the full-strength seawater near the mouth of the estuary, then incubate samples with treatments of additional nitrogen, phosphorus, both, or neither, measuring rates of chlorophyll a growth and determining limiting nutrients.

This study complements data previously collected by others and extends part of the data set, allowing me to compare concentration changes, if any, immediately after the sewer system came online. I (1) estimate the groundwater influx of contaminants to the estuary, (2) create a nutrient profile detailing the distribution of contaminants in the water, and (3) conduct an incubation experiment testing how phytoplankton biomass responds to additions of nitrogen, phosphorus, or both. The nutrient profile offers a different picture of how ammonium, nitrate, and phosphorus concentrations change with depth and salinity, and the nutrient addition bioassay provides new information on the factors limiting phytoplankton growth.

Methods

Groundwater Sampling. To measure nutrient concentrations in the groundwater, I visited each of 12 wells around the perimeter of Little Pond, as well as the stream that discharges water into the estuary (Figure 3). Using a GeoPump ®, I ran well water through a sampling cup into which a Hydrolab Quanta ® sensor array was placed, and I recorded dissolved oxygen (in parts per million and as percent saturation), specific conductivity and salinity, temperature, and pH. I attached to the end of the Geopump ® tubing a 25 mm Swinnex filter containing a GF/F glass fiber filter to fill labeled 60-mL bottles with filtered water. I stored the samples in a cooler with ice.

Nutrient Analysis. Once I returned to the lab, I analyzed each sample for NH4 using the indophenol-blue reaction (modification of Solórzano 1969, Strickland and Parsons 1972), PO4 (modification of Murphy and Riley 1962), and NO3 using a Lachat for Flow Injection Analysis, a procedure based on the cadmium reduction method (Wood et al. 1967). Finally, I tested for total dissolved nitrogen by potassium persulfate oxidation and ran the samples in the Lachat (modification of D’Elia et al. 1976).

Nutrient/Salinity Profile. To build a nutrient profile of the water in Little Pond, I first visited four different stations along the middle of the estuary in a boat (Figure 3). The first station was near the northern end where the stream deposits freshwater into the embayment; the last was closer to the southern end of the estuary. At each station, I used the Hydrolab QuantaTM to take dissolved oxygen, conductivity and salinity, temperature, and pH measurements. Then I used a Geopump ® and Swinnex filters to fill 60-mL and 20-mL sampling bottles with filtered water from varying depths and stored the samples in a cooler with ice.

In the laboratory, I added 60 μL of 5 N hydrochloric acid to each 60-mL sampling bottle and placed both the 60-mL bottles and 20-mL scintillation vials in a refrigerator until I analyzed them for NH4, PO4, and NO3, using the same protocols detailed above. I additionally analyzed samples of non-acidified water from the scintillation vials for conductivity (Radiometer Analystics model CDM 210, temperature corrected to 25 °C), then converted these into salinity measurements using a salinity conversion calculator (Arain 2018).

Surface Water Sampling and Enrichments. I determined the limiting nutrients in Little Pond by adding nitrogen, phosphorus, or both to incubated surface water samples. I visited three different stations along the estuary in a boat, as well as the stream to the north (Figure 3). Based on data from the Hydrolab casts for the nutrient profile, I intended to collect water from water of different salinities of around 7 ‰, 18 ‰, and 24 ‰. At each station, I used a Geopump ® to fill a 20-L carboy with surface water.

Once I returned to the laboratory, I filled 10 bottles each with 850 mL of water from one carboy and repeated for the other three stations for a total of 40 bottles.

I set aside eight bottles total—the first and last bottles I filled from each carboy—and poured a recorded volume of water through 47 mm GF/F filters, plus another recorded volume of the same water through 25 mm GF/F filters. I used the 47mm filters for chlorophyll a (Chl a) analysis (Lorenzen 1967), and the 25 mm filters for particulate organic carbon and nitrogen (POC/N) and total particulate phosphorus (TPP) analysis (Harwood et al. 1969).

I used the remaining 32 bottles for incubation. Within each group of eight bottles from each station, I spiked two samples with 4.25 mL of 16,000 μM ammonium nitrate (“+N”) intended to increase the ambient N concentration by 160 μM. To another two bottles in the group, I added 4.25 mL of 2,000 μM phosphate stock solution (“+P”) to increase the ambient P concentration by 10 μM. I added both of these N and P treatments to another set of two bottles (“+N+P”), leaving the remaining two bottles to serve as controls (“Control”). Finally, to all 32 bottles, I added 5 mL of a mixed inoculum made from 200 mL of water from each carboy. I incubated the bottles at 15°C in a growth chamber set to a daily light/dark cycle of 16 hours at a light level of 410-μE, and 8 hours in the dark, for a total of eight days. To quantify changes in Chl a concentrations day by day, I measured in-vivo fluorescence in a 10-mL sample collected from each bottle with a fluorometer, repeating this measurement every day in the incubation period. I also collected about 20 mL of water from each bottle on the first, fourth, and seventh days, following the nutrient analysis procedure from the groundwater sampling component with those samples. At the end of the incubation period, I filtered between 300 and 400 mL of water from each bottle for Chl a analysis, plus another 100 to 200 mL for each of two filters, one for POC/N analysis and one for TPP analysis.

Results

Groundwater Sampling. Dissolved oxygen in parts per million ranged between 0.46 ppm and 4.84 ppm for all the wells (Table 1, Table 2). Water from the west side, with an average of 1.7 ppm, had more dissolved oxygen compared to water from the east side, with an average of 0.9 ppm. As percent saturation, dissolved oxygen ranged between 4.0% and 54.5%, with the average being 16.4% for the west side and 8.4% for the east side. Conductivity was between 0.023 mS/cm and 46.5 mS/cm. The average in the west was 7.1 mS/cm, smaller than the average in the east, 14.6 mS/cm. Salinity for all the wells were between 0.01 ‰ and 29.91 ‰, with the average in the west, 4.4 ‰, being smaller than the average in the east, 9.1 ‰. The groundwater was slightly acidic in most wells, ranging between 4.27 and 7.53 with an average of 5.79. The average pH of the well water on the west side, 5.7, was only slightly more acidic than that on the east side, 5.9. Finally, the water temperature was between 12.24°C and 19.41°C for all the wells, with an average of 14.0°C. The groundwater on the west side, with an average of 13.7°C, was cooler than the water on the east side, with an average of 14.4°C.

The ammonium concentrations, ranged between 0 μmol/L and 244 μmol/L, with an average of 28 μmol/L among all wells (Figure 4). The highest concentrations were generally on the east side of the estuary, which had an average of 44.1 μmol/L compared to the west side’s 13.2 μmol/L and the stream’s 12.6 μmol/L (Table 3).

The nitrate concentrations ranged between 0 μmol/L and 548 μmol/L, with an average of 88 μmol/L between all wells. The stream had the highest average concentration of 139 μmol/L, followed by the west (108 μmol/L) and east (66 μmol/L) sides. Only nine of the 46 well points had concentrations greater than 200 μmol/L, with six of those points on the west side and three on the east side.

Consolidating the nitrate and ammonium, the mean DIN concentration across all wells was 116 μmol/L, or 121 μmol/L on the west side, 110 μmol/L on the east side, and 152 μmol/L in the stream.

The phosphate (DIP) concentrations were between 0.9 μmol/L and 15.4 μmol/L, the average between all wells being 2.8 μmol/L. The east side had the highest average phosphate concentration at 3.4 μmol/L, followed by the west side (2.3 μmol/L) and the stream (1.7 μmol/L).

The average N:P (DIN:DIP) ratio between all wells was 82; on the west side of the pond only, the ratio was 98; on the east, 66; and in the stream, 92.

Finally, the total dissolved nitrogen concentrations ran between 5 μmol/L and 656 μmol/L between all wells with an average of 145 μmol/L (Figure 5). The stream had the highest average TDN concentration at 181 μmol/L, followed by the western shore (147 μmol/L) and the eastern shore (143 μmol/L).

Compared to similar data collected in 2016 and 2017, DIN concentrations typically followed consistent vertical patterns within each well site (Figure 6). For example, in well LP3, higher concentrations were seen at shallow depths, but lower concentrations were seen at deeper depths for each year. However, there was high variability between different well sites, with some having average concentrations between about 60 and 80 μmol/L (LP1, LP2, LP7, LP10, LP11, LP12), others between 124 and 174 μmol/L (LP3, LP4, LP9), and still others between 214 and 234 μmol/L (LP5, LP6, LP8). Across the years, while some wells showed increases and others showed decreases in nutrient concentrations, there was no consistent trend. For instance, DIN decreases from 2016 to 2017 and 2018 in LP1, LP2, and others. At the same time, DIN increases overtime in LP12, and both rises and falls overtime in LP9 and LP11.

Nutrient/Salinity Profile. Stations 1 and 2 showed strong vertical stratification. The most freshwater estuarine location, station 1, had an average temperature of 7.6 °C; an average conductivity of 32 mS/cm; an average dissolved oxygen content of 9 ppm (87 % saturation); and an average pH 7.49 (Table 4). Salinity was 5.9 ‰ at the surface, increased to 26.5 ‰ between 0.06 m and 0.25 m, then stabilized around 28.3 ‰, and 29.1 ‰ at 0.5 m and 1 m of depth, respectively (Figure 7). The slightly saltier station 2 had an average temperature of 6.9 °C; an average conductivity of 39 mS/cm; an average dissolved oxygen content of 10 ppm (94 % saturation); and an average pH 7.68. Its salinity was 7.8 ‰ at the surface, then increased sharply to 26.4 ‰ at 0.25 m and reached 29.7 ‰ by 1.5 m. In even saltier station 2.5, the average temperature was 6.7 °C; the average conductivity 44 mS/cm; the average dissolved oxygen concentration 9.5 ppm (94 % saturation); and the average pH 7.77. Its salinity was 21.0 ‰ at the surface, then increased to 26.2 ‰ at 0.25 m and 30.0 ‰ at 2 m. Finally, in the saltiest location, station 3, the average temperature was 6.6 °C; the average conductivity 46 mS/cm; the average dissolved oxygen concentration 8.9 ppm (88% saturation); and average pH 7.48. The salinity at the surface was 28.6 ‰, then increased only slightly to 29.6 ‰ by 2 m.

While salinity generally increased with depth, nutrient concentrations generally decreased with depth at each station and were inversely correlated with salinity (Figure 8, R2 = 0.98). At freshwater station 1, DIN concentrations fell from 49 μmol/L at the surface to 8 μmol/L at 1 m. Although the DIP concentration increased slightly from 0.11 μmol/L to 0.22 μmol/L between the surface and 0.06 m deep, below 0.06 m DIP was undetectable. At station 2, the decrease in DIN concentrations was only slightly less dramatic, with a concentration of 43 μmol/L at 0.02 m deep first increasing to 47 μmol/L at 0.06 m, then falling to 7 μmol/L by 2.5 m. DIP concentrations dropped from 0.18 μmol/L at 0.02 m to 0.14 μmol/L at 0.06 m, below which DIP was again undetectable. At station 2.5, the surface DIN concentration of 37 μmol/L decreased to 5 μmol/L by 2 m. At no depth were DIP concentrations high enough to be detected at this station. Lastly, at station 3, DIN concentrations decreased from 10 μmol/L at the surface to 5 μmol/L by 2 m. Once again, no DIP was detectable at any depth at this station.

The N:P ratio of dissolved nutrients was much higher than the Redfield ratio of 16:1 at all sites. Between stations 1 and 2, the N:P ratio was about 470 at the surface, 237 at 0.02 m, and 276 at 0.06 m. On average, the N:P ratio in the surface water was about 315. Below these depths the N:P ratio was indeterminate as a result of no detectable P.

On the plot of salinities against DIN concentrations, the regression line for the estuarine points, 푦 = −2.05푥 + 63 (R2 = 0.98, N = 18), suggests that freshwater at a salinity of 0 ‰ would have a DIN concentration of 63 μmol/L (Figure 8). However, the salinity of the stream water, with a salinity of 0.42 ‰, had a DIN concentration of 152 μmol/L. Similarly, the regression line for the estuarine points suggests the average well water salinity, 6.6 ‰, would align with a DIN concentration of about 77 μmol/L. This is not the case, as the average well water DIN concentration was 115 μmol/L. Nutrient concentrations were lower in the water column than in the groundwater and stream, implying some uptake or removal.

Surface Water Sampling and Enrichments. The bottles containing stream water (salinity = 0 ‰) had average ambient DIN concentrations of 97 μmol/L in both the control and +P treatments. The bottles enriched by the +N or the +N+P treatments contained approximately 228 μmol/L on average, so the treatment added about 130 μmol/L DIN out of the targeted 160 μmol/L (Figure 9, leftmost column of graphs). Average DIN slightly increased in the control treatments, fluctuated in the +N treatments, and generally decreased in the +P and +N+P treatments, but overall not much change was apparent. The average ambient DIP concentrations in the control and +N treatments were both 1.0 μmol/L, and the initial DIP concentrations in the +P and +N+P treatments were both about 14 μmol/L, so that the ambient DIP concentrations were increased by 13 μmol/L, slightly more than the targeted increase of 10 μmol/L. Average DIP in both the control and +N treatments remained low (<1 μmol/L) throughout the experiment, but in both the +P and +N+P treatments decreased from 14 μmol/L to 8 μmol/L, indicating significant uptake.

During the first six days of the incubation, average in-vivo chlorophyll a fluorescence for the control treatments remained fairly stagnant around 0.40 μg/L before picking up after day six, reaching 1.01 μg/L by day eight. The same pattern occurred in the +N treatments, which started at 0.478 μg/L before eventually growing to 1.26 μg/L. Both the +P and +N+P treatments had greater growth, from 0.46 μg/L to 4.21 μg/L in the +P bottle and from 0.45 μg/L to 5.6 μg/L in the +N+P bottles, showing that the addition of P alone may have had a more significant effect on algal growth than the addition of N alone.

The bottles with water from station 1 (salinity = 8 ‰) had an average ambient DIN concentration of 43 μmol/L in the control and +P treatments; since the bottles enriched by +N and +N+P treatments had 166 μmol/L DIN, I added about 120 μmol/L of the targeted 160 μmol/L to the control and +P treatments (Figure 9, second column of graphs). Average DIN decreased in every treatment, especially in the +P treatments (from 43 μmol/L to 1.3 μmol/L) and +N+P treatments (from 168 μmol/L to 114 μmol/L). The ambient DIP concentrations in the control and +N bottles were each about 1.4 μmol/L, and the initial DIP concentrations in the +P and +N+P treatments were about 13 μmol/L, meaning about 12 μmol/L DIP were added via treatment. The control and +N treatments had low initial DIP concentrations (1.0 and 1.9 μmol/L, respectfully), and there seemed to be uptake, as concentrations dropped below 0.5 μmol/L after seven days. The average DIP decreased more so in the +P treatments (from 14 μmol/L to 2 μmol/L) and in the +N+P treatments (from 13 μmol/L to 0.6 μmol/L). Greater reductions in DIN and DIP in the +P and +N+P treatments suggest greater uptake than that in the control and +N treatments.

Over the same incubation period, the average in-vivo chlorophyll a fluorescence in the control treatments grew from 0.41 μg/L, peaking at 3.23 μg/L, and in the +N treatments, 0.38 μg/L to 2.49 μg/L. There was more growth in the +P treatments (from 0.42 μg/L to 14.75 μg/L) and in the +N+P treatments (from 0.38 μg/L to a peak of 14.5 μg/L on day seven), again suggesting greater uptake in the treatments with an addition of P.

The bottles with water from station 2 (salinity = 18 ‰) had ambient DIN concentrations of 35 μmol/L in the control and +P treatments (Figure 9, third column of graphs). The +N and +N+P treatments had initial DIN concentrations of 161 μmol/L, making a 130-μmol/L increase in ambient concentrations in the control and +P treatments. There was some uptake in every treatment on average: DIN did not decrease much in the control, but it did in both the +N and +N+P treatments (from 161 μmol/L to 99 μmol/L) and in the +P treatments (from 35 μmol/L to 1.7 μmol/L). Additionally, the ambient DIP concentrations in the +N and +N+P treatments were each about 0.3 μmol/L, and the initial DIP concentrations in the +P and N+P treatments were 13 μmol/L, so the treatment increased the ambient concentrations by 13 μmol/L P. Average DIP concentrations in the control and +N treatments once again remained low (<1 μmol/L) throughout the incubation. In contrast, there was significant uptake once again in the +P and N+P treatments; average DIP decreased from 13 μmol/L to 2 μmol/L in the +P treatments and from 13 μmol/L to 0.5 μmol/L in the +N+P treatments.

Over the incubation period, the average in-vivo chlorophyll a fluorescence increased in every treatment: in the control treatments, from 0.44 μg/L to 1.95 μg/L; in the +N treatments, from 0.43 μg/L to 1.52 μg/L; in the +P treatments, from 0.41 μg/L to 13.7 μg/L; and in the +N+P treatments, from 0.45 μg/L to a peak of 19.8 μg/L on day six. Treatments with the addition of P, as they have so far, showed more uptake compared to treatments without additional P.

Lastly, the bottles containing water from station 3 (salinity = 27 ‰) had average ambient DIN concentrations of about 13 μmol/L in the control and +P treatments, and initial DIN concentrations of 132 μmol/L in the +N and +N+P treatments, so about 120 μmol/L were added by treatment (Figure 9, fourth column of graphs). Average DIN decreased considerably in all treatments: in the control treatments, from 13 μmol/L to 1.8 μmol/L; in the +N treatments, from 131 μmol/L to 113 μmol/L; in the +P treatments, from 12 μmol/L to 1.5 μmol/L; and in the +N+P treatments, from 133 μmol/L to 48 μmol/L. The average ambient DIP concentrations in the +N and +N+P treatments were both about 1.2 μmol/L, and the average initial DIP concentrations in the +P and +N+P treatments were both about 13 μmol/L, showing the spike added about 12 μmol/L P. Average DIP concentrations in the control treatments remained low (<1 μmol/L) throughout, and in the +N treatments, decreased slightly from 1.4 μmol/L to 0.4 μmol/L. In other treatments, the DIP reductions were greater, from 13 μmol/L to 8 μmol/L in the +P treatments and from 13 μmol/L to 1.6 μmol/L in the +N+P treatments. Once again, significant uptake in treatments with additions of P is apparent.

Overall, the chlorophyll response increased with increasing salinity. The average in-vivo chlorophyll a fluorescence for the control treatments increased from 0.45 μg/L to a peak of 4.12 μg/L on day six. The +N treatments saw an increase from 0.55 μ/L to a peak of 3.26 μ/L on day three. In the +P treatments, fluorescence increased from 0.48 μg/L to a peak of 5.11 μg/L on day six, and in the +N+P treatments, from 0.59 μg/L to a peak of 17.6 μg/L on day five.

Chlorophyll a extracted from particulate material at the end of the experimental incubation showed much higher values than the in-vivo fluorescence. There was a dramatic increase in extractable chlorophyll over the course of the incubation, with a strong response to the +P treatments in the stream and at station 1 and a strong response to the +N+P treatments at all stations. The average initial chlorophyll a concentration of stream water was 0.4 μg/L and grew modestly to only 13 μg/L in the control treatments and to only 14 μg/L in the +N treatments (Figure 10). However, the average grew to 102 μg/L in the +P treatments, and to 115 μg/L in the +N+P treatments. This indicates strong P limitation in the stream water. The average initial extractable chlorophyll a concentration of surface water from station 1 was 0.9 μg/L, and in the control, +N, and +P treatments, it rose to 13 μg/L, 17 μg/L, and 41 μg/L, respectively. But in the +N+P treatments, it increased by two orders of magnitude to 111 μg/L. This indicates strong P limitation and secondary N limitation. Chl a in surface water from station 2 had an initial concentration of 0.9 μg/L; this increased tenfold to 9 μg/L in the control treatments, to 9.5 μg/L in the +N treatments, and to 10 μg/L in the +P treatments. In the +N+P treatments, it rose dramatically to 79 μg/L. This indicates co-limitation by N and P. Finally, Chl a in surface water from station 3 had an initial concentration of 2 μg/L, which remained the same in the control treatments, doubled to about 5.2 μg/L in the +N treatments, decreased slightly to 1.5 μg/L in the +P treatments and increased to 67 μg/L in the +N+P treatments.

Particulate organic carbon (POC), as a measure of C-fixation over the course of the experiment, increased in response to growth of phytoplankton within each bottle during the incubation period. The average initial POC concentration in the stream water was 32 μmol/L (Figure 10). It increased to 117 μmol/L in the control treatments, to 121 μmol/L in the +N treatments, to 244 μmol/L in the +P treatments, and to 312 μmol/L in the +N+P treatments. This suggests a pronounced P limitation. The average initial POC concentration in surface water from station 1 was 74 μmol/L, and it increased over threefold to 246 μmol/L in the control treatments, nearly threefold to 205 μmol/L in the +N treatments, over fivefold to 397 μmol/L in the +P treatments, and eightfold to 599 μmol/L in the +N+P treatments. Again, this suggests P limitation. In bottles with water from station 2, the average initial POC concentration was 69 μmol/L. It rose in nearly the same proportions for every treatment as it did in the station 1 water—almost threefold to 184 μmol/L in the control treatments, less than that to 149 μmol/L in the +N treatments, nearly fivefold to 317 μmol/L in the +P treatments, and nearly eightfold to 508 μmol/L in the +N+P treatments. Once again, there is more P limitation. Lastly, in bottles with surface water from station 3, the initial average POC concentration was 26 μmol/L. It increased to 162 μmol/L in the control treatments, to 134 μmol/L in the +N treatments, to 164 μmol/L in the +P treatments, and higher up to 525 μmol/L in the +N+P treatments. This suggests co-limitation of N and P.

The correlations between Chl a extractions and POC in the stream water, 푦 = 57.9푥 + 1587 (R2 = 0.86), and in the stations altogether, 푦 = 22.9푥 + 869 (R2 = 0.89), at least show that lower Chl a concentrations tend to match with lower POC concentrations, and higher Chl a concentrations tend to match with higher POC concentrations (Figure 11). The slopes are greater than 1, so that there is more POC for every unit of Chl a than there would be if there were a 1:1 ratio between Chl a and POC. The slope in the stream was more than double that in the stations, so perhaps

Particulate organic nitrogen (PON) in the control and +N treatments were generally comparable, but increased more often in the +P and +N+P treatments. The average initial PON concentration in the stream water was 1.7 μmol/L. This increased to 11 μmol/L in the control treatments, slightly higher to 13 μmol/L in the +N treatments, higher to 37 μmol/L in the +P treatments, and highest to 48 μmol/L in the +N+P treatments (Figure 12). The average initial PON concentration in surface water from station 1 was 7 μmol/L. This rose to 22 μmol/L and 24 μmol/L in the control and N treatments, respectively. It rose higher to 44 μmol/L in the +P treatments, and higher still to 83 μmol/L in the +N+P treatments. In surface water from station 2, the initial average PON concentration was 6 μmol/L, and it increased to 18 μmol/L in the control treatments and to 19 μmol/L in the +N treatments. There was a greater increase to 36 μmol/L in the +P treatments, and an even greater increase to 77 μmol/L in the +N+P treatments. Finally, the initial average PON concentration in the surface water of station 3 was 4 μmol/L. This increased to 13 μmol/L in the control treatments, to 18 μmol/L in the +N treatments, slightly less to 14 μmol/L in the +P treatments, and much more significantly to 88 μmol/L in the +N+P treatments. In most cases, PON increased when P was added, and in all cases, PON increased dramatically when both N and P were added. Although initial total particulate phosphorus (TPP) values were not determined, final concentrations per treatment were still available. Like PON, these concentrations were similar in the control and +N treatments and considerably higher in the +P and +N+P treatments, though +P treatment became more comparable to the control and +N treatment concentrations with increasing salinity. The average TPP concentration in the stream water was 1.2 μmol/L and 1.8 μmol/L in the control and +N treatments, and several times higher, at 7 μmol/L and 8.4 μmol/L, in the +P and +N+P treatments (Figure 12). The average TPP concentration in surface water from station 1 was 2.5 μmol/L and 2.4 μmol/L in the control and +N treatments, and again several times higher, at 11 μmol/L, for both the +P and +N+P treatments. In surface water from station 2, the average TPP concentration was 2.2 μmol/L and 2.6 μmol/L for the control and +N treatments, and 9 μmol/L and 10 μmol/L in the +P and +N+P treatments. Things are a little different in station 3 surface water, though. The average TPP concentrations were 2 μmol/L in both the control and +N treatments, not far from the 4 μmol/L in the +P treatments. The concentration was over double that, at 10 μmol/L, in the +N+P treatments.

Particulate N:P (PON:TPP) ratios suggest how the composition of the algae taking up different nutrients may reflect the uptake of those nutrients. In the stream. These ratios were highest in the control (9), then +N (7), +N+P (6), and +P (5) (Figure 13). In station 1, the highest ratio was in +N (10), followed by the control (9), +N+P (8), and +P (4). In station 2, the highest was in the control (8), followed by +N+P (8), +N (7), and +P (4). Finally, in station 3, the highest was in +N+P (9), followed by +N (8), the control (7), and +P (4). It is sensible that treatments without additions in P would generally have higher ratios, meaning more N relative to P was present.

Finally, the plots comparing particulate and dissolved forms of N and P against each other showed a rather weak correlation with N but a fairly strong correlation with P. The difference between final and initial PON concentrations plotted against the difference between initial and final DIN concentrations had the regression line 푦 = 0.82푥 − 1.7 (R2 = 0.50), so that lower differences in PON tended to accompany lower differences in DIN, and higher differences in PON tended to accompany higher differences in DIN (Figure 14). The treatments also had more PON than expected from a 1:1 ratio between PON and DIN. In plotting TPP concentrations against the difference between initial and final DIP concentrations, the regression line 푦 = 1.3푥 − 2.2 (R2 = 0.94) suggests that lower changes in DIP accompany lower TPP values and higher changes in DIP accompany higher TPP values (Figure 15). The correlation is stronger in this relationship compared to that between PON and DIN, and the treatments may have had more DIP than expected from a 1:1 ratio between TPP and DIP.

Discussion

Groundwater Sampling. While there are high N and P concentrations in the wells in 2018, no clear trend in N concentrations is apparent between 2016, 2017, and 2018. Some wells show consistently decreasing N levels while others show increasing N. Still others increase one year but decrease the next, and vice versa. Therefore, I cannot conclude that the sewer system in the Little Pond watershed has had a significant effect on nutrient concentrations within the estuary just yet. The data collected here adds to the baseline measurements that can be compared with data in the future.

A key takeaway from the groundwater sampling portion of this project is that the N concentrations proportional to the P concentrations at this point in time are much higher than expected by the Redfield Ratio, 16N:1P. This should drive the ecosystem toward P limitation.

Surface Water Nutrient Profile. For estuarine sites with freshwater layers from the stream flowing above a saltwater wedge, like station 1, the sharp increase in salinity with depth and decrease in nutrient concentrations demonstrate intense stratification in the northern end of the estuary. This stratification gradually gives way to more thorough mixing in the south, as shown by station 3’s consistent salinities and nutrient levels down the water column.

Again, N:P ratios are much higher than Redfield’s 16N:1P. Together the high N concentrations in groundwater sampling and nutrient profile results, especially compared to P concentrations, suggest that Little Pond may face a P limitation during this season rather than an N limitation.

The tight linear trend between DIN and salinity in the estuarine water shows conservative mixing; there is not much loss or addition of nutrients along the salinity gradient. Since the water was cold, between 6.5 °C and 8 °C, microbial processes such as coupled nitrification and denitrification may have been slowed, increasing the N:P ratio. (Perhaps the rates of microbial processes may increase in warmer seasons and lead to lower N:P ratios in the water.) However, while the stream deposits nutrient-rich freshwater to Little Pond, the difference between the 63 μmol/L predicted by the regression line for the estuarine water samples and the 152 μmol/L determined explicitly in the stream suggests nearly 100 μmol/L DIN is removed, perhaps by uptake, between the time it moves through the stream and the time it flows into the estuary. The same appears to occur between the groundwater wells and the estuary, where some removal makes up the difference between actual groundwater DIN concentrations and that predicted with the regression line for estuarine points.

Surface Water Sampling and Enrichments. DIN and DIP uptake in most bottles and nearly simultaneous rise in Chl a in-vivo fluorescence during the incubations suggests that the dissolved nutrients were taken up by phytoplankton and converted into biomass. For the stream, Station 1, and Station 2 bottles, increased DIN and DIP uptake and subsequent Chl a growth in the +P and +N+P bottles, compared to more modest nutrient uptake and Chl a growth in the +N and control bottles, demonstrates a primary P-limitation suggested by the groundwater, stream water, and nutrient profile investigations discussed above. For the Station 3 bottles, increased nutrient uptake and Chl a growth specifically in the +N+P bottles, compared to more modest nutrient uptake and Chl a growth in bottles with other treatments, show that the most marine site is more co-limited by both P and N.

The pattern of P limitation in low-salinity water and N limitation in high-salinity water has been found in other experiments. For instance, Caraco et al. (1987) looked at nutrient limitation in a series of coastal ponds in Falmouth, Massachusetts, not including Little Pond. These ponds also had densely populated watersheds and high nutrient loading from septic tanks, and enriched surface water from these locations in incubated samples showed that all systems were nutrient limited both in July and in October. Although there was little correlation between salinity and phytoplankton fluorescence in +N+P treatments, such a correlation existed in +N and +P treatments separately. Less saline treatments at less than 6.5 ppt tended to be P limited, more saline treatments at 31 ppt tended to be N limited, and intermediately saline treatments between 6.8 and 30 ppt were on the borderline between N and P limitation. Compared to their experiment, the P limitation in Little Pond was extended between 0 and 18 ppt, and the N and P colimitation appeared around 27 ppt, or at least in water no less saline than 18 ppt. Similarly, Doering et al. (1995) in their own bioassays along an estuarine salinity gradient found a P limitation at 0, 5, and 10 ppt salinities, and an N limitation at 25 ppt salinity. Again, Little Pond’s salinity range of P limitation overlapped and possibly extended beyond that in Doering et al.’s experiment. N limitation in Little Pond was not determined at any salinity below 27 ppt, unlike the N limitation seen in Doering et al.’s experiment starting at 25 ppt.

On the other hand, the conclusion on primary P limitation in Little Pond may conflict with those of previous experiments. Ryther and Dunstan (1971) reported N limitation for algal growth in the Long Island bays of New York. Likewise, Vince and Valiela (1973) found a primary N limitation with a secondary phosphate limitation in Vineyard Sound, as the addition of N rather than P increased chlorophyll concentrations relative to the control. The Long Island bays and Vineyard Sound are more marine than the Little Pond stations sampled here, so that the N limitation at higher salinities was consistent across studies. But since Vince and Valiela (1973) found greater increases in chlorophyll concentrations when both N and P were added to incubated water samples, the major difference between this experiment and theirs was the relative influence of one nutrient over the other with regard to algal growth. Whether N or P is the primary limiting nutrient in coastal bodies of water has been debated at length, and oftentimes N limitation garners more support than P limitation. Still, Howarth and Marino (2006) claim that in some cases, it may come down to a particular system’s characteristics, including presence of N-fixing primary producers, salinity, microbial processes, and nutrient loading and availability. Slow growth rates, paired with grazing mortality by benthic animals and zooplankton, can inhibit the growth of N-fixing cyanobacteria; a dearth of N-fixing primary producers, seen especially in saline sites, can maintain an N deficiency relative to P, leading to N limitation in these systems. P limitation, however, may occur in other estuaries with high N inputs and tight control of P inputs, as seen in the North Sea estuaries of the Netherlands. Ultimately, this experiment helps reveal the nutrient limitations of Little Pond specifically, and can add to existing knowledge on other estuaries in general, too.

With regards to the particulate data (Chl a extractions, POC, PON, and TPP) in Little Pond, concentrations were generally highest across all stations for the +N+P treatments, and more times than not, the +P treatment followed. The difference between +N and +P particulate data also seemed to narrow when moving from stream data to Station 3 data. This may reflect the primary P limitation in less saline sites in Little Pond, and an increasing colimitation that may eventually give way to an N limitation in more saline sites. Additionally, although their slopes were not exactly equal to one, the regression lines in Figure 14 (slope = 0.82) and Figure 15 (slope = 1.3) were fairly close to one, suggesting that the dissolved N and P were largely the source of the N and P present in the growing algae.

Along the salinity gradient, greater nutrient uptake and greater in-vivo fluorescence demonstrate that more saline water generally accompanies greater phytoplankton growth. Since the ambient DIN concentration generally decreased with greater salinity, perhaps phytoplankton in nutrient-rich freshwater already had enough N and did not need to take up much or any more, unlike the phytoplankton in saltier water than had greater N uptake.

As far as these samples show, most of Little Pond may not be N-limited as generally expected from estuaries. Such intense N loading from anthropogenic sources may have caused Little Pond to deviate from the norm for estuaries. Further studies of the ecosystem and its nutrient dynamics can further clarify whether or not this is the case.

Acknowledgements

I would like to thank my mentor, Ken Foreman, for guiding me through this project. I also extend thanks to Rich McHorney, Audrey Rowe, Sarah Messenger, Nick Patel, Anne Giblin, the Ecosystems Center staff at the Marine Biological Laboratory, and my fellow Semester in Environmental Science students. References Arain, H.M. Salinity conversion calculator. https://www.hamzasreef.com/Contents/Calculators/SalinityConversion.php.

Benning-Shorb, J. 2017. Nutrient budget in a eutrophic Cape Cod estuary. Semester in Environmental Science, Marine Biological Laboratory.

Cape Cod Commission. 2017. Watershed Report: Upper Cape – Little Pond.

Caraco, N., A. Tamse, O. Boutros, and I. Valiela. 1987. Nutrient limitation of phytoplankton growth in brackish coastal ponds. Can. J. Fish. Aquat. Sci. 44: 473–476.

Diaz, R.J. and Rosenberg, R. 2008. Spreading dead zones and consequences for marine ecosystems. Science, 321: 926–929.

Earisman, M. and Foreman, K. 2016. A pre-sewering baseline study of nutrient imports and exports from the Little Pond Estuary. Semester in Environmental Science, Marine Biological Laboratory.

Elser, J.J., Andersen, T., Baron, J., Bergström, A.K., Jansson, M., Kyle, M., Nydick, K.R., Steger, L., and Hessen, D.O. 2012. Shifts in Lake N:P Stoichiometry and Nutrient Limitation Driven by Atmospheric Nitrogen Deposition. Science, 326: 835–837.

Harwood, J.E., van Steenderen, R.A., and Kühn, A.L. 1969. A comparison of some methods for total phosphate analyses. Water Research in Permagon Press, 3: 425–532.

Howarth, R.W. and Marino, R. 2006. Nitrogen as the limiting nutrient for eutrophication in coastal marine ecosystems: Evolving views over three decades. Limnol. Oceanogr., 51: 364–376.

Howes, B.L., Eichner, E.M., Saminym, R.I., Kelley, S.W., and Schlezinger, D.R. 2006. 13ed Watershed-Embayment Model to Determine Critical Nitrogen Loading Thresholds for the Little Pond System, Falmouth, MA. SMAST/DEP Massachusetts Estuaries Project, Massachusetts Department of Environmental Protection, Boston, MA.

D’Elia, C. F. and Steudler, P. A. 1977. Determination of total nitrogen in aqueous samples using persulfate digestion. Limnology and Oceanography, 22: 760–764.

Doering, P. H., C. A. Oviatt, B. L. Nowicki, E. G. Klos, and L. W. Reed. 1995. Phosphorus and nitrogen limitation of primary production in a simulated estuarine gradient. Marine Ecology Progress Series, 124: 271–287.

Lorenzen, C.J. 1967. Determination of chlorophyll and pheo-pigments: spectrophotometric equations. Limnology and Oceanography, 12: 343–346.

Massachusetts Department of Environmental Protection. 2003. The Massachusetts Estuary Project: Embayment Restoration and Guidance for Implementation Strategies. Moss, B. et al. 2011. Allied attack: climate change and eutrophication. Inland Waters, 1: 101– 105.

Murphy, J. and Riley, J. P. 1962. A modified single solution method for the determination of phosphate in natural waters. Analytica Chimica Acta, 27: 31–36.

Nixon, S.W. 2009. Eutrophication and the macroscope. Hydrobiologia, 629: 5–19.

Rabalais, N. N., Turner, R. E., Díaz, R. J., and Justic, D. 2009. Global change and eutrophication of coastal waters. ICES Journal of Marine Science, 66: 1528–1537.

Ryther, John H., and William M. Dunstan. 1971. Nitrogen, Phosphorus, and Eutrophication in the Coastal Marine Environment. Science, New Series, 171: 1008–1013.

Solórzano, L. 1969. Determination of ammonia in natural waters by phenol hypochlorite method. Limnology and Oceanography, 14: 799–800.

Strickland, J.D.H. and T.R. Parsons. 1972. A practical handbook of Seawater Analysis. Ottawa, Fisheries Research Board of Canada 2nd. Ed.

Vince, S., and I. Valiela. 1973. The Effects of Ammonium and Phosphate Enrichments on Chlorophyll a, Pigment Ratio and Species Composition of Phytoplankton of Vineyard Sound. Marine Biology, 19: 69–73.

Wood, E.D., Armstrong, F.A.J., and Richards, F.A. 1967. Determination of nitrate in sea water by cadmium-copper reduction to nitrate. Journal of the Marine Biological Association of the United Kingdom, 47(01): 23–31. Figures and Tables

Figure 1. Little Pond locus and Sewer Service Area (SSA).

Figure 2. Nitrogen loading to Little Pond from controllable sources.

Figure 3. Groundwater (left, yellow dots), nutrient profile (right, blue and orange dots), and surface water and enrichment (right, yellow-ringed dots) sampling sites from Little Pond.

Table 1. Hydrolab QuantaTM readings for well water on the west side of Little Pond.

Table 2. Hydrolab QuantaTM readings for well water on the east side of Little Pond.

Figure 4. DIN (green) and DIP (purple) concentrations by depth for western (left) and eastern (right) wells.

Table 3. DIN, DIP, and TDN concentrations in freshwater stream.

Figure 5. TDN concentrations for the northern (top) and southern (bottom) wells by depth.

Figure 6. DIN concentrations determined in 2016 (blue), 2017 (green), and 2018 (purple, my data) by depth for western (left) and eastern (right) wells.

Table 4. Hydrolab QuantaTM readings for water at varying depths mid-estuary.

Figure 7. Salinity, DIN concentrations, and DIP concentrations by depth and station mid- estuary, plus average N:P ratios at different depths.

Figure 8. DIN concentration patterns by salinity mid-estuary, compared with DIN concentrations in the freshwater stream and the average DIN concentration (+/- standard deviation) for all groundwater wells.

Figure 9. Average DIN, DIP, and Chl a in-vivo fluorescence values over the eight-day incubation period for control (blue), +N (green), +P (purple), and +N+P (red) treatment bottles.

Figure 10. Average Chl a extraction concentrations (top) and POC concentrations (bottom) for each sampling site by treatment: initial (pink), control (blue), +N (green), +P (purple), and +N+P (red).

Figure 11. Chl a extracted concentrations against POC concentrations.

Figure 12. Average PON concentrations (top) and TPP concentrations (bottom) for each sampling site by treatment: initial (pink), control (blue), +N (green), +P (purple), and +N+P (red).

Figure 13. N:P (POC:TPP) ratios for each sampling site by treatment: control (blue), +N (green), +P (purple), and +N+P (red). Figure 14. Change in PON concentrations (Day 8 – Day 0) against change in DIN concentration (Day 0 – Day 7).

Figure 15. TPP concentrations against change in DIP concentration (Day 0 – Day 7).

Figure 1. Locus of Little Pond (red indicator, left panel) and outline of Sewer Service area (SSA) in watershed of Little Pond showing outlines of individual parcels (right panel). More than 1400 homes in the SSA will eventually be connected to the sewer and their wastewater will be diverted from the pond.

Figure 2. Nitrogen loading to Little Pond from septic systems, all controllable sources excluding septic systems, and total (including natural sources). Target total loading is indicated with a dashed arrow. Adapted from Howes et al. (2006).

Figure 3. Sampling sites in Little Pond. (Left) The yellow dots are locations of wells for groundwater samples. (Right) The blue dots pinpoint locations in the estuary where water for the nutrient profile was collected, and the orange dot pinpoints the stream. All dots ringed in yellow are locations where carboys of water collected for surface water incubations.

Table 1. Hydrolab QuantaTM readings for well water on the west side of Little Pond. DO is dissolved oxygen.

Well Location DO DO Conductivity Salinity Temperature and Depth (ppm) (% sat) (mS/cm) (‰) pH (°C) LP1 2.2 m 0.75 6.6 0.197 0.09 6.39 13.66 LP1 4.3 m 0.55 4.8 0.295 0.14 6.22 12.88 LP1 6.0 m 0.46 4.0 0.339 0.16 5.77 12.30 LP2 2.1 m 0.68 6.3 0.075 0.04 5.77 13.97 LP2 4.2 m 2.82 26.1 0.124 0.06 5.61 13.32 LP2 6.0 m 4.84 44.2 0.333 0.16 5.44 12.90 LP2 7.9 m 0.80 7.2 0.402 0.19 5.45 12.60 LP3 2.0 m 4.82 46.3 0.229 0.11 5.24 13.76 LP3 4.15 m 1.30 12.1 0.318 0.15 5.25 12.80 LP3 6.0 m 0.94 9.1 0.304 0.14 6.10 12.55 LP3 7.9 m 0.79 7.8 19.9 11.67 4.73 12.67 LP4 2.1 m 0.77 7.4 0.234 0.11 6.44 14.08 LP4 3.25 m 0.48 4.6 0.200 0.09 5.96 13.96 LP4 3.9 m 5.63 54.5 0.202 0.10 5.93 13.75 LP4 6.0 m 0.73 7.8 41.8 26.29 4.82 13.15 LP5 1.2 m 3.00 29.0 0.237 0.11 5.53 13.96 LP5 2.35 m 1.90 18.5 0.215 0.10 5.48 14.32 LP5 3.9 m 0.62 6.8 41.3 26.00 5.45 14.28 LP5 6.0 m 0.71 7.8 44.9 28.56 5.45 15.03 LP6 1.1 m 3.52 34.5 0.193 0.09 6.84 14.92 LP6 2.3 m 3.07 30.6 0.253 0.12 6.32 15.27 LP6 4.35 m 0.81 7.6 0.154 0.07 5.84 14.64 LP6 6.0 m 0.52 4.8 0.301 0.14 5.49 13.82 LP6 7.9 m 0.54 5.3 17.9 10.48 4.89 13.39

Table 2. Hydrolab QuantaTM readings for well water on the east side of Little Pond. DO is dissolved oxygen.

Well Location DO DO Conductivity Salinity Temperature and Depth (ppm) (% sat) (mS/cm) (‰) pH (°C) LP7 2.2 m 1.90 17.5 0.023 0.01 6.01 12.76 LP7 4.0 m 1.29 11.7 0.440 0.21 5.98 12.45 LP7 6.0 m 1.24 11.3 0.396 0.19 6.01 12.25 LP7 7.9 m 1.57 14.4 0.403 0.19 6.06 12.24 LP8 2.4 m 0.97 8.8 0.145 0.07 5.75 13.86 LP8 3.3 m 0.94 8.9 0.288 0.14 5.14 13.61 LP8 4.3 m 0.84 7.9 3.65 1.91 4.53 13.32 LP8 6.0 m 0.85 8.5 22.9 13.66 4.27 13.09 LP8 7.7 m 0.89 9.0 28.8 17.45 4.38 12.90 LP9 1.8 m 0.83 8.4 0.262 0.13 5.45 16.69 LP9 3.8 m 0.54 6.7 39.9 25.21 5.67 18.70 LP9 7.5 m 0.51 6.5 46.5 29.91 5.97 19.41 LP10 2.0 m 0.93 9.0 0.495 0.23 6.67 14.81 LP10 3.4m 0.65 6.3 0.625 0.30 6.45 15.26 LP10 4.6 m 0.60 6.8 43.6 27.62 5.82 14.54 LP11 2.0 m 0.68 6.5 0.140 0.06 7.53 14.78 LP11 3.8 m 0.64 6.7 15.6 8.99 6.28 15.11 LP11 7.2 m 0.55 6.3 44.0 27.87 6.18 13.91 LP12 2.4 m 0.67 6.5 0.335 0.15 7.00 14.38 LP12 3.8 m 0.46 4.4 0.169 0.08 6.48 14.78 LP12 5.2 m 0.65 6.8 29.8 18.22 6.08 14.12 LP12 7.9 m 0.61 6.9 43.4 27.36 6.38 13.67

Figure 4. DIN (green) and DIP (purple) concentrations by depth for western (left) and eastern (right) wells. Note the scales for nutrient concentrations between DIN and DIP, as well as the mean concentrations displayed at the top for wells on the western and eastern shores.

Table 3. DIN, DIP, and TDN concentrations in freshwater stream.

DIN DIP TDN Water Sample (μmol/L) (μmol/L) (μmol/L) Stream A 152.1 1.7 187.1 Stream B 151.9 1.6 174.4 Average for Stream 152 1.7 181

Figure 5. TDN concentrations for the northern (top) and southern (bottom) wells by depth. Note that different wells have different X-axis scales.

Figure 6. DIN concentrations determined in 2016 (blue), 2017 (green), and 2018 (purple, my data) by depth for western (left) and eastern (right) wells. Note that different wells have different X-axis scales. Table 4. Hydrolab QuantaTM readings for water at varying depths in Stations 1, 2, 2.5, and 3 mid-estuary.

Depth Temperature Conductivity DO DO Salinity Location (m) (°C) (mS/cm) (ppm) (% sat) pH (‰) 0 8.49 10.54 9.2 81.8 7.69 5.9 0.06 8.59 12.9 9.25 84 7.54 7 0.1 7.92 37.6 8.34 81.7 7.19 13 Station 1 0.25 6.96 42.1 9.01 88.2 7.34 26.5 0.5 6.65 45 9.27 92 7.52 28.29 1 6.72 46.2 9.21 91.7 7.66 29.05 0 7.95 13.79 10.42 92.2 7.64 7.81 0.25 6.71 42.5 9.38 91.9 7.47 26.44 Station 2 0.5 6.65 44.4 9.9 97.9 7.71 28.08 1 6.64 46.1 9.76 96.5 7.8 28.91 1.5 6.77 47.3 9.24 92.5 7.8 29.74 0 7.08 36.1 9.38 90 7.67 21 0.25 6.71 43.8 9.33 91 7.67 26.23 Station 0.5 6.67 44.3 9.4 92.8 7.73 27.67 2.5 1 6.64 45.4 9.54 94.3 7.81 28.29 1.5 6.58 46.7 9.56 95.1 7.85 29.32 2 6.65 47.7 9.93 99.1 7.91 30.02 0 6.61 45 8.5 85.2 6.6 28.56 0.25 6.61 45.7 8.81 87.2 7.4 28.49 0.5 6.61 45.7 8.94 88.5 7.62 28.63 Station 3 1 6.58 46.5 8.93 88.8 7.7 29.1 1.5 6.55 46.9 9.04 90.1 7.74 29.46 2 6.56 47.1 9.09 90.5 7.79 29.6

Figure 7. Salinity, DIN concentrations, and DIP concentrations by depth and station mid- estuary, plus average N:P ratios at different depths.

Figure 8. DIN concentration patterns by salinity mid-estuary, compared with DIN concentrations in the freshwater stream and the average DIN concentration (+/- standard deviation) for all groundwater wells.

Figure 9. Average DIN, DIP, and Chl a in-vivo fluorescence values over the eight-day incubation period for control (blue), +N (green), +P (purple), and +N+P (red) treatment bottles.

Figure 10. Average Chl a extraction concentrations (top) and POC concentrations (bottom) for each sampling site by treatment: initial (pink), control (blue), +N (green), +P (purple), and +N+P (red). Error bars are standard deviation.

Figure 11. Chl a extracted concentrations against POC concentrations, separated into one regression line (orange) for the stream and another regression line (blue) for stations 1, 2, and 3 combined.

Figure 12. Average PON concentrations (top) and TPP concentrations (bottom) for each sampling site by treatment: initial (pink), control (blue), +N (green), +P (purple), and +N+P (red).

Figure 13. N:P (POC:TPP) ratios for each sampling site by treatment: control (blue), +N (green), +P (purple), and +N+P (red).

Figure 14. Change in PON concentrations (Day 8 – Day 0) against change in DIN concentration (Day 0 – Day 7).

Figure 15. TPP concentrations against change in DIP concentration (Day 0 – Day 7).