Impact of crab bioturbation on benthic flux and nitrogen dynamics of Southwest Atlantic intertidal marshes and mudflats

Estuarine, Coastal and Shelf Science 92 (2011) 629e638

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Impact of crab bioturbation on benthic flux and nitrogen dynamics of Southwest Atlantic intertidal marshes and mudflats Eugenia Fanjul a, b, *, María C. Bazterrica a, b, Mauricio Escapa a, b, María A. Grela b, c, Oscar Iribarne a, b a

Departamento de Biología, FCEyN, Universidad Nacional de Mar del Plata, CC 573 Correo Central B7600WAG, Mar del Plata, Argentina Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Argentina c Departamento de Química, FCEyN, Universidad Nacional de Mar del Plata, Funes 3350, Mar del Plata, Argentina b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 27 November 2009 Accepted 2 March 2011 Available online 31 March 2011

The effect of the SW Atlantic intertidal burrowing crab Neohelice (Chasmagnathus) granulata on benthic metabolism, benthic flux, and benthic N cycling processes was investigated through field experiments and in situ benthic chambers incubations. Our experimental results show that the presence and activity of N. granulata and its burrows may affect the direction and magnitude of nutrient benthic fluxes. Bioturbation enhanced ammonium efflux at mudflats, and influx at marshes. The flux of nitrate toward the sediment was stimulated by crabs at light and dark conditions in marshes, but only under light exposure in mudflats. Crab bioturbation stimulated benthic metabolism, N mineralization, nitrification and denitrification potentials in both sites. Crabs seem to have contrasting effects on dissolved inorganic nitrogen (DIN) availability between marshes and mudflats, as reflected on benthic DIN flux. This different effect on DIN availability and also the possible different effects of crabs on N2-fixers organisms could explain the opposite N2 fixation pattern found for both habitats, since crabs promoted N2 fixation in marshes, but diminished its rate in mudflats. Thus, the results obtained here through manipulative field experiments using benthic chambers suggest that macrofauna may influence the N benthic cycle and DIN fluxes in estuarine sediments. Besides, these macrofauna effects could be context-dependent, being many of them opposite between mudflats and marshes. We concluded that the above mentioned effects and the bioturbationemacrophytes interaction may be affecting the dissolved nutrient exportation from marshes to open waters. Ó 2011 Elsevier Ltd. All rights reserved.

Keywords: bioturbation benthic flux nitrogen cycle intertidal sediment biogeochemistry Argentina

1. Introduction Intertidal sediments are productive and dynamic boundary systems (Nixon, 1986, 1992; Alongi, 1998). Because the water column is shallow, benthic processes tend to be more important than pelagic processes (Sundbäck and McGlathery, 2005). In areas with soft bottoms, the most important biogeochemical processes take place in sediments, where the density of microorganisms in the top layers is several orders of magnitude higher than in the water column (Sundbäck and McGlathery, 2005). The most important link between benthic and pelagic food webs is the deposition of organic matter (OM), its remineralization in the top sediment, and its subsequent return to the water as recycled nutrients that support pelagic primary production (Alongi, 1998).

* Corresponding author. Departamento de Biología, FCEyN, Universidad Nacional de Mar del Plata, CC 573 Correo Central B7600WAG, Mar del Plata, Argentina. E-mail address: [email protected] (E. Fanjul). 0272-7714/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.ecss.2011.03.002

Thus, there is a close, local coupling among deposition, remineralization and benthic flux in these systems (Jahnke et al., 2003). Nitrogen (N) availability plays an important role in supporting primary production in coastal environments (Herbert, 1999). N remineralization (i.e., OM breakdown that produces ammonium) is among the main sources of bioavailable N in these systems (Herbert, 1999; McGlathery et al., 2004). Several microbiologically-mediated processes determine the speciation of recycled N, and consequently, its availability for new primary production (Risgaard-Petersen, 2004). Processes such as nitrification (i.e., the conversion of ammonium to nitrate) and denitrification (i.e., the conversion of nitrate to gaseous nitrogen) are of fundamental importance for N availability in estuarine systems (Risgaard-Petersen, 2004) because the coupling between them implies N out of the system. Most of the bioavailable N is recycled several times between autotrophic and heterotrophic organisms, and rates of N input via N2 fixation (i.e., biological reduction of N2 to ammonia) and N output via denitrification are at least one order of magnitude smaller than internal recycling rates (Canfield et al., 2005).

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Benthic flux largely determines the link between the sediment and the water column (Alongi, 1998). This exchange is due to physical processes of molecular diffusion (a slow process of random molecular motion; Crank, 1979), water advection and sediment resuspension, and complex, biologically mediated processes such as bioturbation and bioirrigation (Burdige, 2006). In most bioturbated sediments, bioturbation is the main transport mechanism (Burdige, 2006). Bioturbating infauna have an impact on benthic flux and on N cycling processes (e.g., polychaeta: Kristensen, 1988; Nielsen et al., 2004; crustacea: Hughes et al., 2000; Webb and Eyre, 2004; bivalvia: Mayer et al., 1995; Michaud et al., 2006). It is widely accepted that bioturbation enhances benthic flux up to an order of magnitude when compared with diffusional flux (Burdige, 2006), and that bioturbators may enhance plant N uptake (e.g., Daleo et al., 2007; Holdredge et al., 2010) and nitrification and denitrification rates (e.g., Mayer et al., 1995; Nielsen et al., 2004). However, the actual bioturbation effect on benthic flux and N cycle depends on the interaction of multiple factors such as bioturbator type (e.g., feeding and burrowing modes, irrigation, mucus secretion; Pelegri and Blackburn, 1995; Michaud et al., 2005), OM reactivity (Kristensen, 2001), interactions between micro- and meio-fauna (Solan and Wigham, 2005). Plant root metabolism and high N demand for primary production may also modify the benthic N cycle in marshes. High rates of N2 fixation occurred in the rhizosphere, coupled to the photosynthetic activity of the plants via the exudation of labile organic carbon (LOC) by the root system (see Welsh, 2000 for review). Moreover, nitrification and denitrification rates depend on the relative influence of oxygen input and on the competition for DIN (i.e., ammonium and nitrite plus nitrate) among plants and nitrifying and denitrifying bacteria (Welsh et al., 2000). Additionally, marsh vegetation creates a three-dimensional mosaic of fluctuating redox surfaces through oxygen injection and root exudations (Holmer et al., 2002; Koretsky et al., 2005) that, for example, enhances the remineralization pathways of the more refractory dissolved organic carbon (Gribsholt and Kristensen, 2002). Thus, the enhanced transport and spatial heterogeneity caused by the interaction between marsh vegetation and bioturbators may cause the coexistence of multiple OM remineralization pathways with diverse effects on benthic flux (Gribsholt and Kristensen, 2002; Koretsky et al., 2005). Intertidal soft sediments along the SW Atlantic coast (from southern Brazil, 23
S to the northern Argentinean Patagonia, San Matías Gulf 41
S) are inhabited by the burrowing crab Neohelice (Chasmagnathus) granulata (e.g., Boschi, 1964; Iribarne et al., 2003). This deposit-feeding crab generates extensive burrowing beds with up to 60 burrows per m2, each of which generally extends from 25 to 100 cm deep and up to 17 cm wide (Iribarne et al., 1997; Botto and Iribarne, 2000). Crab beds can cover 80% of the intertidal areas of this region (see Iribarne et al., 2005; Botto et al., 2006). Burrows affect the particle movement and near-bottom fluid dynamics, working as passive traps of sediments and OM (Botto et al., 2005, 2006), detritus (Iribarne et al., 2000; Botto et al., 2006) and polluting agents (Menone et al., 2004). This species has direct and indirect ecological effects on an ecosystem scale, producing large perturbations on benthic community structure due to sediment changes (Botto and Iribarne, 1999; Iribarne et al., 2005). Unlike other burrowing species (see Kristensen and Kostka, 2005), N. granulata does not actively irrigate its burrows, displaying low irrigation coefficient values due to passive irrigation (e.g., by tidal currents; Fanjul et al., 2007). The excavation of semi-permanent burrows increases sediment surface area, drainage, sediment redox potential (Fanjul et al., 2007), and sediment oxygen concentration (Daleo et al., 2007). Through their excavation activities (daily removal of up to 2.5 kg dry-sediment m2 in the marsh and 6 kg

dry-sediment m2 in the mudflat; Iribarne et al., 1997; Botto and Iribarne, 2000; Escapa et al., 2008), crabs continually rework at least the top 10 cm of the intertidal sediment (Fanjul et al., 2007), leading to a high bioturbation coefficient (250 cm2 yr1; Fanjul et al., 2007). Burrows also reduce the particulate OM exportation from the marshes towards the estuarine waters (Botto and Iribarne, 2000; Botto et al., 2006; Gutiérrez et al., 2006). Although the ecological effects of N. granulata at different scales are well known, there is currently little information about their effects on sediment biogeochemistry (but see Fanjul et al., 2007). Considering the extensive N. granulata crab beds in the SW Atlantic intertidals, their impact in sediment dynamics, and their effect on pore water chemistry, we can anticipate that crabs also influence microbial processes and nutrient flows to neighboring systems. The aim of this work was to determine whether the burrowing crab N. granulata affects benthic metabolism, benthic N cycle processes, and the benthic flux of inorganic nutrients across the wateresediment interface. We hypothesized that N. granulata affects benthic metabolism and N cycling not only by direct mechanisms (i.e., through bioturbation and other activities in their burrows), but also by indirect mechanisms (i.e., by increasing sediment exchange area due to the burrow wall area addition). To test this hypothesis, we conducted field experiments manipulating bioturbation intensity and determined benthic metabolism, benthic N flux, nitrification, denitrification, N2 fixation, and N mineralization by using in situ benthic chambers and laboratory incubation techniques. 2. Material and methods 2.1. Study area The study was conducted near the mouth of the San Clemente tidal creek (eastern part of the Bahía Samborombón, 36
220 S, 56
450 W, Argentina; Fig. 1A,B), a natural protected area affected by low amplitude (up to 1.4 m) semidiurnal tides. Marshes are composed by Spartina alterniflora (at the low zone) and S. densiflora (at the high zone; Isacch et al., 2006). 2.2. Experimental design We conducted a field experiment to evaluate the effect of N. granulata on benthic metabolism and on the flux of dissolved substances across the wateresediment interface. This experiment manipulated bioturbation intensity in three treatments: (1) inclusion of adult N. granulata in artificial burrows (C þ B treatment); (2) unoccupied artificial burrows (B treatment); and (3) exclusion of crabs and burrows used as a “control” treatment (NBC). Plastic mesh boxes (mesh opening 0.75 cm; 0.4 m height, 0.5 m diameter) with a cover of the same mesh (treatment NBC) or without cover (treatments B and C þ B) were used to delimitate each experimental plot. A total of 72 experimental plots were deployed in the field, consisting of 36 boxes in S. alterniflora marsh and 36 boxes in the adjacent mudflat, thus reaching 12 boxes for each treatment and site. Experimental plots were deployed in natural sediments originally not inhabited by crabs and without burrows. Burrow holes (3.5 cm diameter; 40 cm deep) were artificially created at C þ B and B treatments until natural burrow densities for these sites was achieved (mudflat: 29 burrows m2, marsh: 32 burrows m2; E. Fanjul, unpublished data). Each hole was shaped by excavating soil with a PVC core until reaching a depth of 40 cm. This methodology has been used in recent studies performed in similar systems, in which crabs quickly colonized the excavated holes (see Iribarne et al., 2005; Fanjul et al., 2007; Thomas and Blum, 2010). Four months before the beginning of in situ measurements (see

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Fig. 1. Map of study area (A and B); and a close view of benthic chamber construction (C) and its deployment into a field experimental plot (D). Spatial distribution of chambers (into the experimental plots) in Spartina alterniflora marsh during their deployment at flooding tide (E) and during incubation at flooded period (F); and (G) chamber incubation in mudflats during flooded tide. White arrows in (F) and (G) indicate chamber location. Photo credits: M. C. Bazterrica (C, G); M. Escapa (E, F); E. Fanjul (D).

sections 2.3e2.5), one adult crab (between 30 and 40 mm carapace width) per hole was added in C þ B and B treatments. These crabs maintained and often expanded their artificially created burrows. The experimental plots were periodically monitored to ensure that crabs were alive and active in C þ B and B treatments and to check that crabs did not colonize NBC treatments (i.e., fully closed boxes). After two months of inclusion, the crabs inhabiting boxes corresponding to treatment B were removed, and plastic mesh covers were attached to those boxes to inhibit further colonization. Using this procedure, we obtained the configuration for treatment B (i.e., unoccupied crab burrows). After two months, flexible benthic chambers were installed at each experimental plot to measure benthic metabolism, benthic flux, denitrification, and N2 fixation rates (see Fig. 1CeG). All chambers in treatments C þ B and B were placed over a burrow entrance, thus achieving burrow density at the chamber-enclosed sediment equivalent to the natural burrow density (chamber area: 0.0314 m2; 31.8 burrows m2 inside chambers). Detailed benthic chamber construction and usage is described in the following section. 2.3. Benthic flux across the wateresediment interface Benthic flux was determined in all treatments (see section 2.2) using in situ flexible benthic chambers (Fig. 1C). Each chamber was

constructed following Asmus et al. (1998) by mounting a flexible transparent PVC film on top of a PVC pipe (20 cm diameter, 15 cm height). A PVC ring (20 cm diameter, 1.5 cm height) with a transparent lid was attached to the top of the flexible film (see Fig. 1C). The lid contained two septa to sample water inside the chambers using syringes and to take direct measurements with needle sensors. Chambers were inserted 12 cm into the sediment of each experimental plot during tide flooding (see Fig. 1D,E). Special care was taken to avoid trapping air bubbles within the chambers. These flexible chambers transfer the wave action and external water movement into them (Asmus et al., 1998), so it was not necessary to artificially mix the water in the chambers during the incubation period. Incubations were performed during the flooded periods in mudflats and marshes. During all incubations, the water level was between 30 and 100 cm above the chamber lids (Fig. 1F and G). Incubations were conducted both in light and dark conditions (n ¼ 6 each treatment). Dark chambers were shielded from light by wrapping them with opaque aluminum foil. At the end of the incubation, the volume of water inside each chamber was determined by measuring chamber height (i.e., water height) with a ruler. For chambers in C þ B and B treatments, burrow volume was measured by adding a measured quantity of water into each burrow, and this volume was added to each chamber volume. For flux calculations, chambers were considered to have an area of 0.0314 m2 for all treatments.

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Water samples (50 ml) were taken from each benthic chamber 10 min and 3 h after the beginning of incubation in both mudflat and marsh plots. Additionally, water temperature and dissolved O2 concentration were monitored during each sampling period. Water temperature was measured with a hand thermometer, and dissolved O2 was measured in situ by inserting needle sensors (Unisense A/S) through the septa in the chamber lid. Water samples (50 ml) designated for nutrient analysis were filtered in situ (Whatman GF/F) and kept at 20
C until analysis. Salinity, pH and  2 ammonium ðNHþ 4 Þ, nitrate ðNO3 Þ, and sulfate ðSO4 Þ concentrations were also measured. Salinity was measured by refractometry, NHþ 4 was measured by the blue-indophenol method (Solórzano, 2 1969), and NO 3 and SO4 were analyzed by ion chromatography. Flux rates at each chamber were calculated from the differences in solute concentration between time-sequential samples and referred to the enclosed water volume, the surface area of the sediment enclosed, and the duration of the incubation period. The null hypothesis of no crab and light exposure effect on the benthic flux of each substance was evaluated independently for each substance and site (i.e., mudflats and marshes) using two-way ANOVAs (Zar, 1999). We used appropriate a posteriori multiple comparisons tests to evaluate statistical differences among groups (Underwood, 1997). 2.4. N2 fixation Replicated (n ¼ 6) chambers for each treatment (NBC, B and C þ B; in both dark and light conditions) were used to evaluate the effect of crabs and burrows on N2 fixation rates. The N2 fixation rate was measured during incubations using an acetylene (C2H2) reduction assay, as previously described (Stewart et al., 1967). Multiple injections of C2H2-saturated distilled water were performed to achieve a final 10% saturation inside the chambers. Water samples (3 ml) for N2 fixation determination were taken with a syringe 0, 60 and 120 min after C2H2 injection and transferred into evacuated, 12-ml gastight vials containing 2 ml of ammoniacal silver nitrate solution to precipitate the surplus C2H2 (David et al., 1980). Ethylene (C2H4) concentrations in vials were determined by gas chromatography (GC) with a flame ionization detector and a Porapak Q column at 40
C. Ultra high-purity N2 was used as a carrier. N2 fixation was estimated by assuming the theoretical 3:1 ratio between produced C2H4 and fixed N2 (Capone, 1988). The null hypothesis of no crab and light exposure effect on N2 fixation rates was evaluated for each site using two-way ANOVAs (Zar, 1999). The null hypothesis of no crab and site (mudflats and marshes) effect on N2 fixation was evaluated using two-way ANOVA using pooled light and dark data (Zar, 1999). 2.5. Denitrification Denitrification was measured using the acetylene inhibition method. This procedure has frequently been employed to measure denitrification in the past few decades (Groffman et al., 2006). However, it is being used progressively less due to some of its limitations, which can lead to underestimation of denitrification (e.g., inhibition of nitrification). However, in systems as wetlands, this method may be superior to other techniques (Groffman et al., 2006). Moreover, this method (and denitrification potential measurements) is valid for comparisons among treatments and for evaluation of controlling factors (Groffman et al., 2006). 2.5.1. In situ rates In situ denitrification rates were measured in benthic chambers in marsh and mudflat experimental plots (both light and dark conditions; ambient water nitrate concentration 5.50 mmol L1).

Table 1 Statistical summary of two-way ANOVAs of effect of bioturbation and light-dark conditions on benthic fluxes at both sites (mudflats and marshes). Source

Mudflat df

O2 fluxa,b Treatment Light condition Treatment  light condition Error a NHþ 4 flux Treatment Light condition Treatment  light condition Error

NO 3

Marsh MS

F

MS

F

2 1 2

0.1732 0.0277 0.0043

83.5* 13.4* 2.1

1.652 2.722 0.60

432.1* 711.9* 156.8*

30

0.0021

2 1 2

1.7613 0.8338 0.1536

30

0.0676

2 1 2

0.0001 0.0003 0.0001

30

0.0000

0.0038 26.1* 12.3* 2.3

72253.1 565096.7 298946.7

5.8* 45.4* 24.0*

12458.6

a

flux Treatment Light condition Treatment  light condition Error

14.0* 55.0* 18.0*

2072323 316176 74469

160.8* 24.5* 5.8*

12889

SO2 4

flux Treatment Light condition Treatment  light condition Error

2 1 2

5234930 1768494 44138040

30

9818293

0.53 0.18 4.5*

1547352 34782 62914

168.7* 3.8 6.9*

9172

*Indicates p < 0.05. a Log-transformed data (mudflat). b Log-transformed data (marsh).

Nitrate (500 mmol L1), chloramphenicol (0.1 mg L1), and acetylene (10% saturation) were added into the water retained in the chambers. Samples (10 ml) for nitrous oxide (N2O) measurements were collected 0, 30, and 210 min after C2H2 injections. All N2O concentrations were measured in situ by injecting the samples in an evacuated mini-chamber (10 ml) equipped with a N2O microsensor (Unisense AS) and by GC (thermal conductivity detector, Porapak Q column, and ultra high-purity He as carrier). The slope of a lineal regression was used to estimate denitrification rates. 2.5.2. Potential rates When in situ benthic chamber incubations were concluded, sediment cores (20 cm diameter; 5 cm deep) were extracted from each experimental plot (n ¼ 6 each treatment) and then weighed and homogenized at the laboratory. To determine the maximum denitrification potential, a subsample (w15 g) was incubated in the presence of C2H2 (10% saturation). Incubations were accomplished in the dark at ambient temperature (20
C) following the denitrification enzyme activity assay (Groffman et al., 1999). N2O concentration in the headspace was measured 2 and 24 h after the beginning of the incubation by GC as described above. The difference in N2O concentration between initial and final samples was related to the incubated sediment area and the incubation time to estimate the denitrification potential rates. The null hypothesis of no crab and sites effect on denitrification potential was evaluated using two-way ANOVA (Zar, 1999).

2.6. N mineralization and nitrification potential Sediment samples were collected from the same cores used to measure denitrification potential. Samples were homogenized with a spatula and incubated following the short-term laboratory incubation protocol described by Robertson et al. (1999). Gross N

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lowest O2 uptake was in the NBC treatments, with no differences between light and dark incubation. In marshes, the highest O2 uptake was found in C þ B treatments, and NBC treatments exhibited the lowest O2 uptake rates (Fig. 2B). O2 uptake in C þ B and B treatments was higher in light than in dark incubation, and no difference was observed in NBC treatments (see Table 1, Fig. 2B). Ammonium benthic flux was different among treatments and between light and dark incubations in mudflats (Table 1; Fig. 2C). In this site, NHþ 4 fluxes were from the sediment to the water column, except for B and NBC dark incubations that showed a flux towards the sediment (Fig. 2C). The highest flux toward the water column was in C þ B treatments, and the lowest occurred in NBC light treatments (Fig. 2C). In contrast, in marshes NHþ 4 flux showed an interaction effect between treatment and light-dark condition factors (Table 1). NHþ 4 flux was towards the marsh sediments, except for B and C þ B dark treatments, in which there was a NHþ 4 flux toward the water column (Fig. 2D). The highest NHþ 4 flux towards the sediments was in the C þ B light treatment, and the lowest was at NBC light and dark treatments. Nitrate flux showed an interaction effect between treatment and light-dark condition factors for mudflat and marsh sites

mineralization rate was calculated as the increment of DIN, considering incubation time and sediment area. This rate represents the amount of DIN available for either plant N uptake or DIN loss (e.g., benthic flux). The nitrification rate was calculated as the difference between incubated and initial nitrate concentrations according to time and sediment area. The null hypothesis of no crab and site effect on these processes was evaluated using two-way ANOVAs (Zar, 1999). 3. Results 3.1. Benthic flux across the wateresediment interface Salinity and pH did not differ between time-sequential samples at any light condition, experimental treatment or site (i.e., mudflats and marshes). In mudflats, O2 uptake was different among treatments and between light and dark conditions (two-way ANOVA; Table 1). A posteriori comparisons indicated that the highest gross rates of O2 uptake occurred in C þ B treatments, and for this treatment, O2 uptake was higher in dark than in light incubations (Fig. 2A). The

A

B

C

D

E

F

G

H

Fig. 2. Benthic flux of oxygen (A and B), ammonium (C and D), nitrate (E and F), and sulfate (G and H) at light and dark for experimental treatments C þ B, B and NBC in mudflats and marshes, respectively. Here and thereafter, a negative flux rate indicates that the flux was from the water column into the sediment. Values are presented as mean  1SD. Different letters represent significant differences between treatments (LSD test; a ¼ 0.05).

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(Table 1). In mudflats, light-exposed C þ B treatments showed the highest NO 3 uptake, and the lowest uptake was registered in NBC treatments. In contrast, NO 3 flux in dark incubations did not vary among treatments and had similar values to those in the lightincubated NBC treatment (Fig. 2E). In marshes, NO 3 flux was towards the sediment, except for the dark-incubated NBC treatment, which showed a small flux towards the water column. NO 3 uptake was higher in C þ B than in B treatments, and NBC light treatment showed the lowest NO 3 uptake (Fig. 2F). Sulfate flux showed an interaction between factors for both sites (Table 1). All fluxes in the mudflats were towards the sediment, and there were no differences among the light-incubated treatments (Fig. 2G). For dark incubations, the lowest flux towards the sediments was in the C þ B treatments, and the highest was in the NBC (Fig. 2G). In marsh sites, NBC treatments (both light and dark) showed small SO2 4 flux toward the water column, whereas in B and C þ B treatments flux was towards the sediment (Fig. 2H). The flux in C þ B was higher in light than in dark incubations, and these fluxes were both higher than in B treatments. When comparing O2 uptake between marsh and mudflat sites, irrespective of the dark or light condition (i.e., pooled light and dark data), there was an interaction effect between treatment and site factors (two-way ANOVA; Treatment  Site: F2, 66 ¼ 4.58, p < 0.05). The highest O2 uptake was in marsh and mudflat C þ B treatments, whereas the lowest was in both NBC treatments. Furthermore, bioturbation enhanced O2 uptake in the marsh 20-fold, while in mudflats, this enhancement was only 2-fold (Fig. 3). In addition, the benthic flux of DIN (dark and light pooled data) showed an interaction effect between treatment and site (Treatment  Site: F2, 66 ¼ 4.87, p < 0.05). Post-hoc tests showed that in mudflats, bioturbation had no effect on DIN flux, whereas it enhanced sediment DIN uptake in marshes (Fig. 3). 3.2. N2 fixation rates N2 fixation rates in mudflat sediments showed differences among treatments and between light and dark conditions, with no

Fig. 3. Oxygen uptake and DIN benthic flux for C þ B, B and NBC experimental treatments in mudflat and marsh sites for light and dark pooled data. Values are mean  1SE.

Fig. 4. N2 fixation rates (mean  1SD) for light and dark experimental treatments.

interaction effect between these factors (two-way ANOVA; Treatment: F2, 30 ¼ 40.79, p < 0.05; light-dark: F2, 30 ¼ 8.73, p < 0.05; Fig. 4). Post-hoc comparisons indicated that in mudflats, C þ B treatments (in light and dark) presented the lowest N2 fixation rates, whereas the highest N2 fixation rate was in the NBC light treatments (Fig. 4). For this last treatment, N2 fixation was higher at light than at dark conditions. Conversely, in marshes, there were differences among treatments (Treatment: F2, 30 ¼ 14.01, p < 0.05), but no differences between light and dark conditions. The highest rates were found in C þ B treatments, and NBC treatments showed the lowest N2 fixation rates (Fig. 4). Irrespective of light condition, N2 fixation rates showed an interaction effect between site and treatment factors (dark-light pooled data; Treatment  Site: F2, 66 ¼ 37.84, p < 0.05). In marshes, the highest fixation rate was in the C þ B treatments, and the lowest rate was in the NBC treatments. The opposite pattern was observed in mudflats (Fig. 5).

Fig. 5. N2 fixation rates comparing mudflat and marsh experimental treatments for light and dark pooled data. Values are presented as mean  1SD.

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3.3. Denitrification rate In situ denitrification rates were lower than the detection limit of the method used (1 mmol N2 m2 h1) in all treatments and incubation conditions. Denitrification potential rates were higher in the marsh than in the mudflat sediments (Table 2, Fig. 6). Denitrification potentials were 2 and 3 times higher in C þ B and B than in NBC treatments for marsh and mudflat sediments, respectively (Fig. 6). 3.4. N mineralization and nitrification potentials N mineralization potential rates showed an interaction effect between treatment and site factors (Table 2). The highest N mineralization potential was in C þ B treatments, and it was higher in marsh than in mudflat sediments. The lowest N mineralization rate was in the NBC treatment. Moreover, there was no difference in B and NBC treatments between marshes and mudflats (Fig. 6). The rate of nitrification potential also showed an interaction effect between treatments and sites (Table 2). Nitrification potential rates were lower in mudflats than in marshes. In marshes, the rates of C þ B treatments were higher than those calculated for B treatments, and NBC treatment showed the lowest nitrification potential rate. However, the nitrification rate in NBC in marshes and in C þ B treatments in mudflats was similar. The lowest rates were in B and NBC mudflat treatments (Fig. 6). 4. Discussion Few in situ experimental studies have directly determined the biogeochemical effects of bioturbators and their actual impact on the benthic flux and N-cycle processes (but see Webb and Eyre, 2004). Through field experiments manipulating crab and burrow densities, we showed that the dominant SW Atlantic intertidal crab N. granulata may affect the benthic flux, and N-cycling processes in both marsh and mudflat sediments. The effects of bioturbation activities were contrasting between marshes and mudflats for several processes (see Fig. 7). Bioturbation by crabs stimulated benthic metabolism and the potential rates of total N mineralization, nitrification, and denitrification potentials in both mudflats and marshes. N2 fixation was stimulated in the marshes but was inhibited by bioturbation in the mudflats. In turn, crabs stimulated DIN efflux from mudflat sediments, but in the marshes, crabs favored DIN sediment uptake (i.e., flux towards the sediment). In Table 2 Statistical summary of two-way ANOVAs of the effect of bioturbation and sites (mudflats and marshes) on denitrification, N mineralization, and nitrification potential rates. Source

df

Denitrification potential Treatment Site Treatment  site Error

2 1 2 30

N mineralization potential Treatment Site Treatment  site Error

2 1 2 30

16628397 6454182 5131868 112426

147.9* 57.4* 45.7*

Nitrification potential Treatment Site Treatment  site Error

2 1 2 30

4948201 15513354 7096881 49868

99.2* 311.1* 142.3*

*Indicates p < 0.05.

MS

F 2140.8 7458.2 87.48 88.99

24.1* 83.8* 1.0

Fig. 6. Denitrification, total N mineralization, and nitrification potential rates (mean  1SD) for experimental treatments in mudflat and marsh sediments.

situ experiments showed that the presence and activity of N. granulata had more intense effects than just the presence of its burrows for almost all evaluated processes. Whereas unoccupied burrow treatments showed how burrow structure affects biogeochemical processes, occupied burrow treatments showed the effects of exchange area addition (i.e., presence of burrow structure) and effects due to changes in sediment geochemistry. Bioturbation effects on sedimentary characteristics, such as enhanced sedimentcolumn transport processes (i.e., OM, electron acceptors, metabolites transport, see Fanjul et al., 2007), changes in OM quality through the sediment column, or changes on microbial assemblages, may explain these changes on sediment geochemistry. Potential rates of nitrification reported here were on the same order of magnitude as the actual rates reported for other estuarine environments (see Henriksen and Kemp, 1988; Canfield et al., 2005 for reviews), except for bioturbated marsh sediments, where rates were about 10-fold higher. In marshes, bioturbation enhanced the potential rates of total N mineralization and nitrification 8-fold, whereas in mudflats, this enhancement was 2-fold and 4-fold, respectively. These increases in potential rates reflect the effect of bioturbation in microbial populations (see Kristensen, 1988; Hamersley and Howes, 2005), but do not necessarily reflect the

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Fig. 7. Brief scheme of bioturbation effects on N-cycling processes in mudflats (gray plus/minus symbols) and marshes (black plus/minus symbols). Larger plus/minus symbols represent a larger effect of bioturbation stimulating/inhibiting the indicated processes. Dashed arrows showed processes not quantified in this work but that may be affected by crabs. Abbreviations: DNRA (dissimilatory nitrate reduction to ammonium); LOC (labile organic carbon); MPB (mictophytobenthos); OM (organic matter).

actual rates of these processes. For example, our laboratory incubations indicated that about 75% of the remineralized N in marshes was nitrified, and only 25% of the remineralized N was being nitrified in mudflats. These last values are consistent with those commonly reported in the literature for mudflats (between 10 and 65%; e.g., Henriksen and Kemp, 1988; York et al., 2010). However, the high proportion of recycled NHþ 4 oxidized by nitrification in marsh sediments may not reflect the processes actually occurring in this bioturbated sediment. This high proportion of nitrification could be due to the fact that all of the recycled NHþ 4 is accumulated during laboratory incubations and is available to nitrifiers, because þ processes such as NHþ 4 -plant uptake or NH4 release by sediments do not occur during incubations. However, these last processes were also stimulated by bioturbation, and may diminish the actual fraction of NHþ 4 oxidized by nitrification. The extended oxic/anoxic interface created by burrow structures may directly increase denitrification by enhancing NO 3 influx, and increase it indirectly by enhancing nitrification (e.g., Pelegri et al., 1994; Nielsen et al., 2004). Several studies have reported that most bioturbators enhance the denitrification rate between 1.5 and 6-fold (e.g., Pelegri and Blackburn, 1995; Canfield et al., 2005). Our results showed that crabs and burrows increased the rate of denitrification potential by 2 and 3-fold in marshes and mudflats, although in situ rates of denitrification were very low at both sites (below 1 mmol m2 h1; see Herbert, 1999; Hamersley and Howes, 2005 for reviews). This undetectable actual denitrification rate may be a result of low carbon content in these sediments (1.2% at mudflats, and 2.0% at marshes; E. Fanjul unpublished results), although further experiments will be necessary to understand the actual role of N. granulata on denitrification and coupled nitrification-denitrification. Solute flux between the sediment and water column plays an important role in estuarine nutrient cycles, and benthic N flux can be used as a net balance of the individual processes involved in sediment N turnover (Blackburn and Henriksen, 1983; Herbert, 1999). In mudflat sediments, the enhancement of NHþ 4 production generated by bioturbation (by increased OM remineralization and probably also by dissimilatory nitrate reduction to ammonium, DNRA) exceeded the increased N uptake by microphytobenthos and nitrifying bacteria, as suggested by benthic NHþ 4 efflux (McGlathery et al., 2004; Hamersley and Howes, 2005). The NHþ 4 release reported here

for non-bioturbated sediments was similar to those reported from analogous systems (e.g., Asmus et al., 1998; Papaspyrou et al., 2007; York et al., 2010), although bioturbated sediments showed high fluxes (up to 6-fold higher than other highly bioturbated environments; Hughes et al., 2000; Papaspyrou et al., 2007). In contrast with our results found in mudflats, bioturbation enhanced NHþ 4 uptake in light-exposed marsh sediments by 3e4fold. This result probably reflects an enhancement in the actual rates of nitrification (as occurred with nitrification potentials; our results), but may be mostly due to the stimulation of NHþ 4 uptake by primary producers (see light-dark oxygen flux results; Pedersen et al., 2004; Daleo et al., 2007). However, during darkness bioturbation shifted the NHþ 4 uptake (in non-bioturbated sediments) to NHþ 4 release in bioturbated marsh sediments, indicating a positive balance between the production and consumption of NHþ 4 . This NHþ 4 release occurred at even higher rates than those reported for other S. alterniflora marshes (e.g., Andersen, 1986; Whiting et al., 1989). Bioturbation also enhanced NO 3 uptake and, although the rate of denitrification potential was enhanced 2-fold by bioturbation, it may not be the effect of denitrification-related NO 3 consumption because the actual denitrification rates seem to be 2 1  very low (below 1 mmol N2 m h ). This high NO3 uptake could be due to uptake by primary producers (because we cannot entirely reject this process even in dark conditions, Rysgaard et al., 1993; Lee and Dunton, 1999), but also may be due to the stimulation of DNRA rates. Indeed, recent studies have shown that DNRA (which includes NO 3 reduction coupled to sulfide oxidation) may be an important process in intertidal sediments (An and Gardner, 2002; benthic flux for dark sediBurdige, 2006). Our results of SO2 4 reduction (and sulfide ments, suggest a stimulation of SO2 4 production; e.g., Aller and Yingst, 1978). Thus, the high NHþ 4 release 2 and high NO 3 and SO4 uptake that occurred in darkness in bioturbated marsh sediments could be related to the coupling between DNRA and sulfide oxidation (Porubsky et al., 2009). N remineralization and sediment metabolism were stimulated by crab activities in mudflats and marshes, and in marshes bioturbation may also promote DIN uptake by enhancing primary production (our results; Montague, 1982; Daleo et al., 2007). This leads to contrasting effects of crabs on benthic DIN flux between marshes and mudflats, with almost no net flux of DIN under any bioturbation condition in mudflats and net DIN uptake stimulated

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by crab activities and burrows in marshes. These results suggest a higher availability of DIN in bioturbated mudflat sediments than in bioturbated marsh sediments. Biological N2 fixation is an energy-demanding process that is unlikely to occur where DIN is available in relatively high concentrations (Fenchel et al., 1998). In mudflats where light is not limited but the availability of LOC is low, N2-fixing photoautotrophs (such as cyanobacteria) have a significant advantage over N2-fixing heterotrophs (Herbert, 1999). In mudflats, bioturbation reduced N2 fixation rates by 3-fold, possibly by enhancing DIN availability but also mainly due to negative effects on benthic cyanobacteria (e.g., limiting their light exposure by burial or by grazing). In spite of this effect, mudflat sediments showed high N2 fixation rates compared to those found in other mudflat sediments (Capone, 1988; Canfield et al., 2005; Bertics et al., 2010). In marshes, N2 fixation by cyanobacterial mats is considered to be less important as an overall source of N than heterotrophic fixation (e.g., by sulfate-reducing bacteria; Fenchel et al., 1998). Several studies suggest that Nlimited plants exude LOC in their rhizosphere, which sets N2 fixation by stimulating sulfate-reducing bacteria (e.g., Fenchel et al., 1998). The activity of N. granulata enhances LOC content (E. Fanjul, unpublished results), reduces DIN availability, and seems to stimulate sulfate-reducing activity (our results; Fanjul et al., 2007). Thus, it is probably by indirectly promoting an array of sulfatereducing bacteria and other heterotrophic N2-fixing bacteria, bioturbation by crabs enhanced N2 fixation rates in marshes by 3-fold. Despite this enhancement of N2 fixation these rates were similar to those previously reported in other similar marshes (Teal et al., 1979; Currin et al., 1996; Canfield et al., 2005). Benthic metabolism, benthic flux, and benthic N cycling are ecosystem functions that regulate many of ecosystem services supplied by marshes and mudflats (e.g., C sequestration, gas regulation, nutrient recycling; Constanza et al., 1997). On a global scale, marshes comprise about 5% of land area; however they support an increasing anthropic impact (Wolanski, 2007). Although SW Atlantic marshes are threatened mainly due to coastal development and land reclamation, several of SW Atlantic marshes are protected areas and remain pristine. These marshes allow evaluating ecosystems functions of natural systems in which bioturbation intensity remain among the highest reported for similar habitats (Fanjul et al., 2007; Wang et al., 2010). In this work we showed how crab bioturbation could regulate many of the valuable ecosystem functions supplied by marshes. Moreover, because the burrowing crab N. granulata and S. alterniflora co-exist across a wide area of SW Atlantic coastal estuaries and bays (Botto et al., 2005; Alberti et al., 2007), their individual and combined effect in benthic biogeochemistry could be a mechanism controlling the estuarine and coastal food web, and the ecosystem services supplied by marshes. For example, our results showed that macrofauna may influence the N benthic cycle and DIN flux in intertidal sediments, and some of its effects were contrasting between mudflats and marshes (see Fig. 7). Indeed, several studies have shown that both bioturbation and rooted macrophytes create an oscillating oxic/anoxic environment (e.g., Aller, 1994; Holmer et al., 2002), where OM remineralization rates are high and the N cycle and benthic metabolism are severely affected. Our results concur with previous studies, but also suggest that the interaction between bioturbation and macrophytes may be affecting dissolved nutrient exportation from marshes to open waters. Acknowledgments This project was supported by grants from the Universidad Nacional de Mar del Plata, the Fundación Antorchas, CONICET, ANPCyT, and PNUD/GEF Patagonia to O. Iribarne. This work is part

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of the Doctoral Thesis of E. Fanjul at the UNMdP. E. Fanjul, M. C. Bazterrica, and M. Escapa were supported by scholarships from CONICET (Argentina). We thank M. Addino for her help on field experiments. We thank I. Valiela and two anonymous reviewers for comments, which substantially improved the manuscript.

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