Carbon and nitrogen isotope ratios of juvenile winter flounder as indicators of inputs to estuarine systems

Marine Pollution Bulletin 101 (2015) 624–631

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Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul

Carbon and nitrogen isotope ratios of juvenile winter flounder as indicators of inputs to estuarine systems Richard J. Pruell ⁎, Bryan K. Taplin U.S. Environmental Protection Agency, Office of Research and Development, National Health and Environmental Effects Research Laboratory, Atlantic Ecology Division, 27 Tarzwell Drive, Narragansett, RI 02882, United States

a r t i c l e

i n f o

Article history: Received 30 September 2015 Received in revised form 13 October 2015 Accepted 16 October 2015 Available online 2 November 2015 Keywords: Stable nitrogen isotopes Stable carbon isotopes Fish Winter flounder Estuaries

a b s t r a c t Stable carbon and nitrogen isotope ratios were measured in young-of-the-year (YOY) winter flounder, Pseudopleuronectes americanus, collected from several Rhode Island, USA estuarine systems. These included three coastal lagoons, an estuarine river and Narragansett Bay. The δ13C trends observed along transects in several systems showed isotopically depleted terrestrial signals in the upper reaches of the estuaries. Significant differences (P b 0.05) in δ15N were observed among all estuarine systems and these differences correlated (P b 0.01) with human population densities in the watersheds. Although Narragansett Bay has a strong north–south gradient in nutrient concentrations this trend was not reflected in flounder δ15N. The northernmost station with the highest nutrient concentrations unexpectedly had significantly lower δ15N values. Depleted δ15N values at this nutrient-rich station may indicate that concentration-dependent fractionation needs to be considered when using nitrogen isotope ratios in biota to monitor anthropogenic nitrogen inputs in systems with high nitrogen loadings. Published by Elsevier Ltd.

1. Introduction Estuarine ecosystems in many areas have been impacted by anthropogenic nitrogen loadings (Howarth et al., 2002; Bricker et al., 2008). As a result, many of these systems show signs of eutrophication such as plankton and macroalgae blooms and depleted oxygen levels (Kemp et al., 2005). In addition, habitat changes such as the loss of submerged aquatic vegetation (Valiela et al., 1992) and salt marsh area (Deegan et al., 2012) impact many estuarine species which use these areas as juvenile habitat. Tracing nutrient distributions in estuarine systems is often complex due to many factors. Nitrogen inputs to estuarine systems have multiple sources which can include river discharges, urban and agricultural runoff, groundwater inputs, atmospheric deposition and wastewater treatment facilities (Castro et al., 2003). Also, estuarine systems generally have complicated flow regimes, and nitrogen biogeochemistry in these systems is exceedingly complex (Herbert, 1999). Therefore, identifying the relative influence of anthropogenic nutrient inputs at different locations can prove challenging. Stable isotope ratios are increasingly used to study ecological systems. Ratios of the naturally occurring 13C and 12C isotopes of carbon (13C/12C or δ13C) in biota can provide information on the relative importance of different sources of carbon to a food web at a particular location

⁎ Corresponding author. E-mail address: [email protected] (R.J. Pruell).

http://dx.doi.org/10.1016/j.marpolbul.2015.10.037 0025-326X/©Published by Elsevier Ltd.

(Gearing et al., 1984; Wada et al., 1990; Prado et al., 2014). Nitrogen isotope ratios (15N/14N or δ15N) can be used to infer the position of species within food webs; however, these ratios can also be used to track anthropogenic nitrogen inputs to aquatic systems. In particular, nitrogen from sewage treatment facilities or in groundwater influenced by septic systems tends to be enriched in 15N relative to 14N (Heaton, 1986; McClelland and Valiela, 1998). Therefore, measurements of δ15N in water, sediments and biota have been used to indicate the relative importance of sewage inputs to aquatic systems and to monitor changes over time at specific locations. Many factors besides anthropogenic inputs can also influence the isotopic ratios of biota including trophic position, life stage and size. Therefore, it is important to limit the variability that is often observed by constraining as many of these variables as possible. Connolly et al. (2013) found that nitrogen isotope ratios indicative of sewage inputs were incorporated equally by all members of the food web. However, Bannon and Roman (2008) reported that consumer species showed lower within-site variability of δ15N values than producer species. Also, numerous studies have successfully used fish tissue δ15N as an indicator of anthropogenic nitrogen inputs (Gaston et al., 2004; Hoffman et al., 2012; Morris et al., 2015). Therefore, we selected a fish species to track anthropogenic nitrogen inputs in several RI estuarine systems. Winter flounder, Pseudopleuronectes americanus, was chosen as our target species because flounder generally spawn in nearshore areas, therefore, they are available at all of our sampling locations and the young-ofthe-year (YOY) shows little movement after settlement during their first summer (Saucerman and Deegan, 1991). To limit variability further

R.J. Pruell, B.K. Taplin / Marine Pollution Bulletin 101 (2015) 624–631

we selected fish of a narrow size range so that the fish were all about the same age, and therefore, likely feeding on similar prey items. The primary goal of this study was to assess the use of carbon and nitrogen isotope ratios in winter flounder as indicators of anthropogenic inputs to estuarine systems. Systems reflecting a wide range

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in the degree of anthropogenic influence were selected and samples were collected over three years to establish the temporal variability of these indicators. Isotope ratios in the winter flounder were then compared with various watershed characteristics of the estuarine systems.

Fig. 1. Sampling locations for YOY winter flounder along the coast of Rhode Island, USA. The sampling sites included three coastal ponds (Ninigret Pond, Green Hill Pond and Pt. Judith Pond), an estuarine river system (Narrow River) and Narragansett Bay. The station identifications are listed in Table 1. Locations and annual effluent flows for the wastewater treatment facilities in the area averaged over 2002–2004 are also shown.

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Table 1 Station salinity and stable isotope measurements of juvenile winter flounder muscle. Fish were collected from three habitat types (U = unvegetated, E = eelgrass and M = macroalgae) along the Rhode Island coastline. δ13C ± S.D. (‰)

δ15N ± S.D. (‰)

Station

Salinity (PSU)

Location

Code

2002

2003

2004

2002

2003

2004

2002

2003

2004

Narragansett Bay Passeonkquis Cove-M Barrington Beach-U Greenwich Bay-M Prudence Island-E Wickford Beach-U East Ferry-E Fox Hill-M Sheffield Cove-U

PCM BBU GBM PIE WBU EFE FHM SCU

19.2 27.9 27.6 30.2 29.4 30.9 31.3 31.1

17.7 26.0 26.4 28.8 29.9 30.8 30.0 30.4

24.8 28.7 28.0 30.5 30.6 31.7 31.4 31.4

−14.8 ± 0.2 −14.1 ± 0.7 −14.4 ± 0.5 −14.7 ± 1.0 −13.4 ± 0.4 −15.4 ± 0.9 −13.6 ± 0.6 −12.5 ± 0.5

−17.3 ± 0.8 −15.8 ± 0.8 −15.9 ± 0.4 −16.1 ± 0.6 −14.2 ± 0.7 −14.9 ± 0.4 −13.7 ± 0.3 −13.6 ± 0.3

−15.7 ± 0.3 −14.4 ± 1.2 −14.3 ± 0.4 −14.4 ± 0.6 −13.5 ± 0.9 −14.4 ± 0.8 −12.2 ± 0.7 −11.7 ± 0.5

13.7 ± 0.4 14.6 ± 0.5 14.3 ± 0.3 14.9 ± 0.7 14.8 ± 0.4 15.0 ± 0.4 13.0 ± 0.6 13.0 ± 0.3

12.7 ± 0.4 14.7 ± 0.5 14.4 ± 0.4 14.7 ± 0.5 15.0 ± 0.3 15.2 ± 0.4 14.5 ± 0.4 14.0 ± 0.4

13.4 ± 0.4 14.1 ± 0.7 14.5 ± 0.2 14.1 ± 0.5 14.9 ± 0.5 14.0 ± 0.5 13.6 ± 0.3 13.3 ± 0.3

Narrow River Narrow River-U Narrow River-E Narrow River-M

NRU NRE NRM

13.4 13.4 28.0

8.5 13.2 24.0

11.0 16.4 29.1

−16.1 ± 0.3 −14.5 ± 0.6 −13.9 ± 0.6

−15.0 ± 0.7 −14.3 ± 0.6 −12.5 ± 0.5

−17.8 ± 1.5 −14.5 ± 1.9 −12.7 ± 0.5

12.9 ± 0.3 13.7 ± 0.2 13.2 ± 0.4

13.1 ± 0.3 13.7 ± 0.4 13.8 ± 0.7

13.0 ± 0.3 13.6 ± 0.4 13.8 ± 0.2

Pt. Judith Pond Pt. Judith-M Pt. Judith-E Pt. Judith-U

PJM PJE PJU

27.3 29.0 29.7

24.8 28.2 29.1

29.1 30.6 31.7

−16.1 ± 0.4 −15.8 ± 0.6 −15.9 ± 0.3

−16.4 ± 0.4 −15.6 ± 0.4 −15.3 ± 0.8

−16.3 ± 0.6 −14.1 ± 0.5 −12.5 ± 0.5

12.8 ± 0.3 12.9 ± 0.4 12.5 ± 0.3

13.1 ± 0.4 13.2 ± 0.4 12.8 ± 0.2

12.8 ± 0.4 13.2 ± 0.4 12.6 ± 0.3

Green Hill Pond Green Hill-M Green Hill-E Green Hill-U

GHM GHE GHU

25.2

20.1

22.8

NA NA −13.4 ± 0.6

−13.8 ± 0.8 −14.7 ± 1.1 −14.7 ± 0.4

−10.8 ± 1.2 NA −12.2 ± 0.7

NA NA 11.4 ± 0.4

11.4 ± 0.5 10.5 ± 0.4 10.7 ± 0.3

11.6 ± 0.5 NA 11.7 ± 0.2

Ninigret Pond Ninigret-M Ninigret-E Ninigret-U

NGM NGE NGU

26.0 30.5 29.3

21.0 30.5 31.0

24.7 29.9 30.8

−14.7 ± 0.6 −12.6 ± 1.0 −14.2 ± 0.8

−13.9 ± 0.6 −13.8 ± 1.0 −15.0 ± 0.7

−13.9 ± 0.8 −11.2 ± 1.3 −12.3 ± 0.9

9.3 ± 0.5 9.2 ± 0.8 11.2 ± 0.2

9.7 ± 0.5 9.1 ± 0.5 10.1 ± 0.7

9.4 ± 0.5 9.0 ± 0.4 11.1 ± 0.9

NA — fish not available.

2. Materials and methods

salinity based on conductivity ratios (practical salinity units). It was calibrated using a YSI 3169 conductivity standard.

2.1. Fish collection 2.2. Sample preparation and isotopic analysis YOY winter flounder were collected from each of 20 stations (Fig. 1; Table 1) along the coast of Rhode Island, USA. These sampling locations included three shallow coastal lagoons, which are locally referred to as ponds (Point Judith Pond, Ninigret Pond and Green Hill Pond), an estuarine river system (Narrow River) and the upper, middle and lower portions of Narragansett Bay (Fig. 1). The coastal lagoons are small, shallow systems that are connected to Block Island Sound by a series of manmade breachways (Lee and Olsen, 1985). Narrow River is a small, shallow estuarine river system (Kelly and Moran, 2002) with a steep salinity gradient (~10–31). Narragansett Bay is a medium sized, well mixed estuary (Pilson, 1985) with a salinity gradient ranging from about 24 to 32. The physical properties of the waterbodies and the characteristics of their associated watersheds are provided in Table 2. Additional details on the locations and physical characteristics of these stations are also included in a previous publication (Pruell et al., 2012). Fig. 1 provides information on the locations and discharge flows (averaged from 2002– 2004) from the wastewater treatment facilities in the watersheds. Fish were collected from three habitat types (eelgrass stands, macroalgae beds and unvegetated areas) within each of the sampling regions. No eelgrass beds are found in the upper areas of Narragansett Bay so only two sites were sampled from that region of the Bay. Also, on several occasions flounder were not available from two of the sites in Green Hill Pond. Ten YOY winter flounder (45 to 65 mm) were collected from each station (Fig. 1; Table 1), resulting in a total of 568 fish collected and analyzed for this study. Fish were collected at low tide from each site using a 100-foot beach seine during June and July of 2002, 2003 and 2004. They were placed on ice upon collection, returned to the laboratory and then frozen. Salinity was measured at each sampling site (Table 1) during fish collections using a YSI Model 85® Handheld meter (YSI Incorporated). This instrument calculated

Fish samples were thawed and measured and the muscle tissues were removed for isotopic analysis. The samples were placed into aluminum pans and placed in an oven at 60 °C until dry. These dry tissues were then ground to a fine powder using a mortar and pestle and stored in glass vials until analyzed. About 1.0 mg of each sample was weighed Table 2 Physical properties and watershed characteristics for the waterbodies studied which included three coastal salt ponds, Narrow River and Narragansett Bay.a Coastal Ponds Ninigret Green Point Narrow Narragansett Hill Judith River Bay Physical characteristics Area of waterbody (km2) Average depth (m) Average salinity (PSU) Watershed characteristics Area of watershed (km2) Agricultural land (%) Barren land (%) Commercial (%) Forest and brush land (%) Industrial (%) Population density (individuals km−2) Residential (%) Urban (%) Wetland (%)

6.4 1.2 28

1.7 0.8 23

7.8 1.8 30

2.4 2.0 10–20

378 8.8 24–33

30.8 2.6 3.1 1.5 65.7 0.0 64

14.2 3.9 1.2 1.0 54.8 0.0 162

80.4 6.6 1.1 1.9 56.7 0.1 227

35.9 4.7 0.1 0.4 62.9 0.2 274

4756 4.7 1.0 2.1 48.8 1.5 395

15.8 2.7 3.8

33.7 0.8 1.7

22.2 2.9 2.9

22.7 2.2 3.5

21.4 4.5 8.7

a Data sources: Rhode Island Geographic Information System (RIGIS, 2013), 2003–2004 data; Massachusetts Office of Geographic Information (MassGIS Land Use, 2005); Lee and Olsen (1985), Keller et al. (1999), Kelly and Moran (2002), Hubeny et al. (2009).

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into a tin boat and analyzed using a GV Instruments IsoPrime (Elementar Americas, Inc.) continuous flow-isotope ratio mass spectrometer (CF-IRMS). Three-point regression lines were used to calibrate all carbon and nitrogen isotope ratio measurements. For carbon this included the international standards, USGS24 (graphite), IAEA-C-6 (sucrose ANU), and DORM-2 (National Research Council of Canada) which was used as in-house standard and all data are reported relative to VPDB. Nitrogen isotope ratios were determined against a calibration line that included IAEA-N-2 (ammonium sulfate), IAEA-N-3 (potassium nitrate) and DORM-2; all data are reported relative to N2 in air. Each of these standards was analyzed at the beginning, middle and end of each instrumental run. A blank sample was also analyzed at the beginning of each sample run to verify that CO2 and N2 backgrounds were low; however, the data were not blank corrected. A sample duplicate was analyzed after every ten samples; the average isotopic difference measured for these duplicate samples was about 0.3‰ for both δ13C and δ15N. 2.3. Statistical analyses Statistical analyses were performed using a series of multi-factor general linear models. When assessing differences between sampling years, analyses performed using a three-factor model, with sampling year and waterbody type as crossed factors, station as a nested effect with waterbody type, and an additional interaction term between sampling year and waterbody type. Due to differences observed from year to year at the station level and not all stations being sampled at all years, comparisons between waterbody types and waterbodies were performed using a separate two-factor model for each year, with waterbody type and station (nested within waterbody type) as model factors. Comparisons across stations were also fit separately to each waterbody type using two-factor models, with year, station and a year-by-station interaction term included in each model. Overall, Ftests of model significance were performed at the alpha = 0.05 significance level. To identify statistically significant pairwise differences between years, waterbodies or stations, Tukey–Kramer and Bonferroni pairwise comparisons were performed with adjustments made based on the number of comparisons of interest to ensure an overall Type I error rate across all comparisons of 0.05. For all models, the assumption of the data following a normal distribution within each year and station was assessed by using both graphical review and testing the model residuals using the Anderson–Darling test, performed at alpha = 0.01. The assumption of homoscedasticity (constant variance across all stations and years) was assessed by plotting the residuals against the fitted model estimates. In all cases assumptions were met. Flounder δ13C trends with station salinity and watershed characteristic comparisons with flounder δ15N values were conducted using Pearson correlations. 3. Results Yearly plots of δ13C versus δ15N for all stations and samples (Fig. 2) show that, overall, similar isotope patterns were observed during the three sampling years (2002–2004). However, carbon isotope ratios were more variable among the three years than δ15N. In particular, the range of δ13C values observed during 2004 was considerably larger than those measured during 2002 and 2003, especially for the Narrow River stations. Carbon and nitrogen isotope ratios did not correlate with fish length at any of the sampling stations. Some significant differences in the δ13C values of winter flounder muscle tissues were observed among stations within Narrow River and Narragansett Bay (Table 1). In Narrow River the lowest δ13C values were at station NRU and highest at NRM. The isotopic ratios at these stations were significantly different (P b 0.05) for all three years (F-statistics for station contrast: 2002: F = 14.2; 2003: F = 17.6; 2004: F = 72.6, two degrees of freedom for each contrast; Bonferroni t-statistics by year: 2002: t = 5.18; 2003: t = 5.70; 2004: t = 11.9). Within

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Narragansett Bay the upper Bay stations (PCM, BBU, PIE and GBM) had the most depleted δ13C values and all three years were significantly (P b 0.05) lower than those observed at SCU which is the southernmost station sampled (F-statistics for station contrast: 2002: F = 19.8; 2003: F = 40.4; 2004: F = 40.5, seven degrees of freedom per contrast; individual Bonferroni t-statistics ranging between 5.28 and 13.7). Also, a significant positive correlation was found between station salinity and δ13C in flounder muscle tissue for all three years along the salinity gradients of Narrow River and Narragansett Bay. For Narrow River highly significant correlations (P b 0.01) were found for all three years (2002, r = 0.635; 2003, r = 0.863; 2004, r = 0.777, 29 degrees of freedom for each year). In Narragansett Bay a significant correlation (P b 0.05) between salinity and flounder δ13C was observed for 2002 (r = 0.275) and highly significant relationships were found for 2003 and 2004 (2003, r = 0.754; 2004, r = 0.634) with 79 degrees of freedom for each year. These trends indicate the influence of riverine and/or terrestrial organic matter along the estuarine salinity gradient of these systems. Each year, winter flounder δ15N values were significantly different among all of the waterbodies sampled (P b 0.05; contrast F-statistics: 2002: F = 560.8; 2003: F = 799.9; 2004: F = 539.0; four degrees of freedom per contrast). Nitrogen isotope ratios increased according to the following sequence: Ninigret Pond b Green Hill Pond b Pt. Judith Pond b Narrow River b Narragansett Bay (Fig. 3) for all three years (F-statistics for pairwise contrasts range between 15.37 and 2503; one degree of freedom per contrast). Also, grouped together, fish from the coastal ponds had the lowest nitrogen isotope ratios followed by Narrow River and then Narragansett Bay (P b 0.05; waterbody type comparison F-statistics: 2002: F = 805.25; 2003: F = 1145.7; 2004: F = 706.6; two degrees of freedom per contrast). No significant differences in δ15N (P N 0.05; F = 2.48; two degrees of freedom) were observed for any of the waterbodies among the three sampling years. Since there were no differences among years, the yearly data for each waterbody were combined and compared with various watershed land use parameters (Table 2). A highly significant correlation (Fig. 4) was observed between watershed population density and δ15N in the flounder collected from the respective waterbodies (r = 0.97; P b 0.01, 567 degrees of freedom). None of the other watershed land use characteristics were significantly correlated with δ15N in the flounder. Although a large gradient in nitrogen concentrations occurs in Narragansett Bay, a relatively small range of δ15N ratios was observed in the flounder tissues throughout the Bay (Table 1). As expected, lower δ15N values were seen at some of the lower-Bay stations (FHM and SCU); however, the remaining sites all had similar nitrogen isotope ratios (Fig. 5) except for the station furthest north in the Bay, PCM, which had lower than expected δ15N ratios for all three years (Table 1). In fact, during 2002 δ15N values in fish from PCM were not statistically different (P N 0.05; Bonferroni t-statistics: FHM: t = 3.59; SCU: t = 3.69) from the lower-Bay stations FHM and SCU, and during 2003 δ15N at PCM was significantly (P b 0.05: Bonferroni t-statistics range between 6.10 (SCU) and 12.2 (EFE) lower than all other Narragansett Bay locations. 4. Discussion A wide range of δ13C values (−19.9 to −9.4‰) were measured in the muscle tissues of YOY winter flounder collected from several estuaries along the Rhode Island coastline. The lowest carbon isotope ratios were observed in the upper reaches of the Narrow River and Narragansett Bay (stations NRU and PCM). Also, there were significant positive correlations between station salinity and flounder δ13C in these systems. This likely reflects the greater influence of terrestrial carbon inputs at these locations due to runoff and sewage inputs (Wada et al., 1987, 1990; Prado et al., 2014). Winter flounder muscle tissues measured in the present study showed a δ13C range of about 4‰ along a north–

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Fig. 2. Plot of δ13C versus δ15N for all YOY winter flounder samples collected in (a) 2002, (b) 2003 and (2004) from the five Rhode Island coastal waterbodies studied.

south transect of Narragansett Bay. Similarly, Gearing et al. (1984) reported an increase of about 3‰ in δ13C ratios of sediments collected along a similar transect of the Bay. Also, the range of δ13C values measured in the flounder in the present study is similar to that reported by Vinagre et al. (2008) for two flatfish species (Solea solea and Solea senegalensis) from an estuary in Portugal. The stations with the highest δ13C values in flounder tissues were in Ninigret Pond and Green Hill Pond and in the lower reaches of Narragansett Bay (stations SCU and FHM). These locations have higher salinities and are influenced to a greater degree by marine carbon sources, which tend to have higher δ13C values. The relative influence of C3 and C4 plants may also be a factor in the δ13C of the biota at particular locations. Marsh grasses, which are C4 plants with relatively enriched δ13C

values (Farquhar et al., 1989; Vinagre et al., 2008), may contribute to the increased δ13C values of the carbon at the base of the food web in Green Hill and Ninigret Ponds and at the Narragansett Bay stations SCU and FHM, which are surrounded by large stands of the salt marsh cordgrass, Spartina alterniflora. Also, Green Hill and Ninigret Ponds contain large beds of eelgrass, Zostera marina, which has elevated δ13C ratios relative to other C3 marine plants (Hemminga and Mateo, 1996). Overall, the trends we noted in winter flounder δ13C values are consistent with those found by Deegan and Garritt (1997) in another New England estuarine system. Considerable variability was observed in the δ13C ratios of the fish muscle tissues among the three years. In particular, the range of values measured during 2004 was much larger than those observed during

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Fig. 3. Mean δ15N values of YOY winter flounder collected from the five Rhode Island coastal waterbodies for each of the three sampling years (2002, 2003 and 2004). Fig. 5. Mean δ15N values of YOY winter flounder collected from the eight sampling stations within Narragansett Bay for each of the three sampling years (2002, 2003 and 2004).

2002 and 2003. Kostecki et al. (2010) found year to year variations in the δ13C values of juvenile common sole (S. solea, L.) muscle tissue collected from estuarine systems and attributed the differences to variations in river flow. During the present study, the lowest river flows were observed during 2004 (Pruell et al., 2012). This reduced terrestrial influence might help to explain why the most enriched δ13C values of the study were observed in some of the flounder tissues from two of the coastal ponds (Ninigret and Green Hill) that year. However, it would not explain the highly depleted ratios observed in Narrow River during 2004. Oczkowski et al. (2014) reported that particulate matter δ13C values increased with nutrient additions in mesocosm experiments and during a spring bloom in Narragansett Bay. Therefore, changes to nutrient inputs to these systems as well as other factors that influence the δ13C values of primary producers or the degree of isotopic fractionation in food webs (Aberle and Malzahn, 2007) may also be factors. Large differences in δ15N were observed in the juvenile winter flounder tissues among the Rhode Island estuarine systems studied. Fish from the coastal ponds had the lowest nitrogen isotope ratios compared with those from Narrow River and then Narragansett Bay. This trend was consistent among the three years studied and no significant differences within a waterbody were observed among years. Among the coastal ponds, Ninigret Pond fish had the lowest δ15N values followed by those from Green Hill and Pt. Judith Ponds. These results are similar to those from caging studies with YOY winter flounder in these waterbodies (Meng et al., 2000, 2001). Some estimates of anthropogenic nitrogen inputs to these estuarine systems are available but comparisons among systems are often difficult. Several studies have measured nitrogen inputs to the Rhode Island

Fig. 4. Correlation between human population density in the watershed and δ15N values of the muscle tissue of YOY winter flounder collected from each of the systems during 2002, 2003 and 2004 (r = 0.97; P b 0.01, 567 degrees of freedom).

coastal ponds and Narrow River (Lee and Olsen, 1985; Kelly and Moran, 2002; Moran et al., 2014); however, input estimates are highly variable among these studies. Also, comparisons between the coastal ponds and Narragansett Bay are problematic because the coastal lagoons and Narrow River receive nutrient inputs mostly from groundwater whereas the largest source of nutrients to Narragansett Bay is from wastewater treatment facilities (Nixon et al., 1995; Vadeboncoeur et al., 2010). Therefore, instead of using estimates of nutrient inputs to each system to compare with fish δ15N we tested numerous land use characteristics of the watersheds against the δ15N measurements of the juvenile flounder from the estuaries systems. The trends in δ15N were found to be highly correlated with human population density of the watersheds and not correlated with any of the other watershed characteristics. Therefore, despite all of the differences in nitrogen input routes and estuarine characteristics, δ15N in fish from these systems correlated closely with the population density of the watershed. This agrees with a study by Brown and Valiela (2001) who reported that the percent contribution of wastewater to total nitrogen inputs increased with watershed population density in a series of small New England estuaries and Bannon and Roman (2008) who observed that, of the watershed characteristics considered, population density was the best predictor of δ15N in estuarine biota including fish. Several other studies have also used δ15N in fish tissues to infer anthropogenic nitrogen distributions. Martinetto et al. (2006) reported that δ15N in numerous fish species reflected land-derived nitrogen sources. Morris et al. (2015) found that percent urbanization was significantly correlated with δ15N in five fish species from several Australian estuaries. Schlacher et al. (2005) and Connolly et al. (2013) reported that δ15N muscle tissue values in various fish species accurately reflected sewage input patterns in several Australian estuarine systems. Also, Radabaugh et al. (2013) demonstrated the use of δ15N isoscapes of fish tissue to track the plume of the Mississippi River in the eastern Gulf of Mexico. In addition, Dierking et al. (2012) measured δ15N in the muscle tissue of common sole (S. solea, L.) and found enriched isotope ratios in fish from Northwestern Mediterranean coastal lagoons compared to fish from open coastal areas. The range of δ15N values that they reported was similar to that of the present study (8.3 to 16.0‰). Higher δ15N values (15 to 19‰) were reported (Vinagre et al., 2008) for two flatfish species (S. solea and S. senegalensis) from the Tagus Estuary in Portugal. More than 80% of the sewage inputs to Narragansett Bay occur in the upper areas of the estuary (Oczkowski et al., 2008). Therefore, there is generally a large (factor of five or greater) north–south gradient in seawater nitrogen concentrations in the Bay (Oviatt et al., 2002; Pruell et al., 2006; Thornber et al., 2008). However, the δ15N trends observed in the fish tissues from the Bay do not show a large decrease with distance down the estuary. One possible reason for this unexpected trend could

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be the differential fractionation of nitrogen isotopes during uptake at the base of the food web (Cifuentes et al., 1989; Hoch et al., 1992) along the nutrient concentration gradient in Narragansett Bay. Thornber et al. (2008) reported that Gracilaria spp. grown in seawater collected from the nutrient gradient in Narragansett Bay also did not show the expected decrease in δ15N with distance down the Bay. They speculated that at stations with high nutrient concentrations such as Gaspee Point (adjacent to our PCM station) Gracilaria preferentially accumulated lighter nitrogen. These findings agree with Montoya et al. (1991) who reported differential δ15N fractionation between NH+ 4 and particulate nitrogen along the nutrient gradient in Chesapeake Bay. Also, Cifuentes et al. (1988) reported fractionation of NH+ 4 by phytoplankton at concentrations above 20 μM in Delaware Bay. In a laboratory studies Pennock − et al. (1996) found isotopic fractionation during NH+ 4 and NO3 accumulation by the marine diatom, Skeletonema costatum at concentrations above 20 μM and Needoba et al. (2003) reported nitrogen isotope fractionation in twelve species of marine phytoplankton grown on high nitrate concentrations. Swart et al. (2014) found that macroalgae also fractionate nitrogen isotopes during the assimilation of nitrate and Viana and Bode (2013) suggested that high nitrogen concentration leads to depleted δ15N values in macroalgae in an estuarine system in Spain. Liu et al. (2013) developed a relationship between ammonium concentration and the degree of nitrogen isotope fractionation by phytoplankton. They reported that the largest extent of isotopic fractionation occurred at ammonium concentrations between about 10 and 60 μM. In a previous study Pruell et al. (2006) measured the con− centrations of NH+ 4 and NO3 as high as 46.7 and 15.9 μM, respectively, at the PCM station. Therefore, the concentrations measured at this site are similar to those that have been shown to cause significant concentration-dependent nitrogen isotope fractionation in phytoplankton and macroalgae. This may lead to the preferential accumulation of a lighter δ15N signature by the base of the food web at this location which is then passed to consumers such as winter flounder. There have been conflicting nitrogen isotope trends in biota reported along the nitrogen concentration gradient in Narragansett Bay. Unlike the present study, Pruell et al. (2006) measured nitrogen isotope ratios in six different species collected at three locations along the nutrient gradient in Narragansett Bay and found that four of the six species sampled did show significant decreases in δ15N with distance down the estuary. Also, Oczkowski et al. (2008) found that macroalgae δ15N values decreased with distance down Narragansett Bay; however, δ15N ratios in hard clams did not. The reason for these differences is not clear but could be related to seasonal changes in δ15N at the base of the food chain. Two studies of New England estuarine systems (Wainright and Fry, 1994; York et al., 2007) have shown that phytoplankton δ15N values were lower during the winter when nitrogen concentrations were elevated and higher during the summer and suggested that these differences may be related to fractionation. Also, Lesutiene et al. (2014) found that isotopic changes at the base of a food web were sometimes reflected in fast-growing secondary consumers such as juvenile fish but not in slow-growing species. Since juvenile winter flounder have rapid nitrogen turnover rates (Bosley et al., 2002) and these YOY flounder grew during the late winter and spring when nitrogen concentrations are highest, these differences may be related to seasonal factors that are not reflected in all species. Chaves (2004) found that mussels collected along the nutrient gradient in Narragansett Bay showed lower than expected δ15N values in the upper region of Narragansett Bay, similar to what we found for the winter flounder. Interestingly, these were small mussels collected during April when nitrogen levels would be elevated in Narragansett Bay (Krumholz, 2012) and mussels, unlike some other bivalves, actively filter (Oczkowski et al., 2008). Therefore, in estuarine systems with high nitrogen loadings seasonal cycles in nitrogen concentrations may influence nitrogen isotope ratios in estuarine food webs; therefore, it is important to consider many factors when using δ15N to infer anthropogenic nutrient exposures in these systems.

5. Conclusions Carbon isotope ratios in YOY winter flounder showed a large range of values among the sampling locations along coastal Rhode Island. At most stations similar isotopic ratios were observed among the three sampling years; however, differences among years were observed at some sites. Overall, the δ13C trends were consistent with the expected relative degree of terrestrial versus marine influence at the various locations. Muscle tissues of YOY winter flounder showed large differences in δ15N among the estuarine waterbodies studied along the coast of Rhode Island. These differences persisted over the three years of the study and were consistent with the extent of anthropogenic influence, indicated by population density in the watershed, in these estuaries. Therefore, these measurements appear useful as a generalized method for monitoring nitrogen inputs to these systems. The nitrogen isotope ratios observed in the highly nutrient enriched upper part of Narragansett Bay, however, suggest that fractionation at the base of the food web due to high nitrogen concentrations may have resulted in isotopic ratios lower than expected at some locations. These findings indicate that δ15N in estuarine biota can be used to infer trends in the relative importance of anthropogenically derived nitrogen in estuarine systems; however, because of complex nitrogen biogeochemistry caution is needed when interpreting these trends in aquatic ecosystems with high anthropogenic nitrogen loadings. Acknowledgments We would like to thank Jennifer Yordy for assistance with the field collection of juvenile winter flounder, Kenneth Miller (Computer Sciences Corp.) for the statistical analyses, Michael Charpentier (Raytheon Corp.) for GIS work and Patricia DeCastro (SRA International, Inc.) for assistance with the graphics. The sample collections were conducted under permit from the Rhode Island Department of Environmental Management. The research for this article was supported by the US EPA, but has not been subject to agency review, and therefore does not necessarily reflect the views of the agency. No official endorsement should be inferred. This is ORD Tracking Number ORD-009821. Mention of trade names or commercial products does not constitute endorsement or recommendation for use. References Aberle, N., Malzahn, A.M., 2007. Interspecific and nutrient-dependent variations in stable isotope fractionation: experimental studies simulating pelagic multitrophic systems. Oecologia 154, 291–303. Bannon, R.O., Roman, C.T., 2008. Using stable isotopes to monitor anthropogenic nitrogen inputs to estuaries. Ecol. Appl. 18, 22–30. Bosley, K.L., Witting, D.A., Chambers, R.C., Wainright, S.C., 2002. Estimating turnover rates of carbon and nitrogen in recently metamorphosed winter flounder Pseudopleuronectes americanus with stable isotopes. Mar. Ecol. Prog. Ser. 236, 233–240. Bricker, S.B., Longstaff, B., Dennison, W., Jones, A., Boicourt, K., Wicks, C., Woerner, J., 2008. Effects of nutrient enrichment in the nation's estuaries: a decade of change. Harmful Algae 8, 21–32. Brown, J.L., Valiela, I., 2001. The ecological effects of urbanization of coastal watersheds: historical increases in nitrogen loads and eutrophication of Waquoit Bay estuaries. Can. J. Fish. Aquat. Sci. 58, 1489–1500. Castro, M.S., Driscoll, C.T., Jordan, T.E., Reay, W.G., Boynton, W.R., 2003. Sources of nitrogen to estuaries in the United States. Estuaries 26, 803–814. Chaves, J.E., 2004. Potential use of 15N to Assess Nitrogen Sources and Fate in Narragansett Bay (Dissertation) University of Rhode Island. Cifuentes, L.A., Sharp, J.H., Fogel, M.L., 1988. Stable carbon and nitrogen isotope biogeochemistry in the Delaware estuary. Limnol. Oceanogr. 33, 1102–1115. Cifuentes, L.A., Fogel, M.L., Pennock, J.R., Sharp, J.H., 1989. Biogeochemical factors that influence the stable nitrogen isotope ratio of dissolved ammonium in the Delaware Estuary. Geochim. Cosmochim. Acta 53, 2713–2721. Connolly, R.M., Gorman, D., Hindell, J.S., Kildea, T.N., Schlacher, T.A., 2013. High congruence of isotope signals in multiple marine taxa. Mar. Pollut. Bull. 71, 152–158. Deegan, L.A., Garritt, R.H., 1997. Evidence for spatial variability in estuarine food webs. Mar. Ecol. Prog. Ser. 147, 31–47. Deegan, L.A., Johnson, D.S., Warren, R.S., Peterson, B.J., Fleeger, J.W., Fagherazzi, S., Wollheim, W.M., 2012. Coastal eutrophication as a driver of salt marsh loss. Nature 490, 388–392.

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