Introduction

The processes that drive ecosystem functioning often involve complex relationships between organisms and their environment. In marine ecosystems, biological perturbation (bioturbation) of sediments is widely acknowledged to affect ecosystem processes by altering fluxes of both energy and matter across strong geochemical gradients. Processes, such as organic matter remineralization, primary production, sediment transport, and nutrient cycling are often measured as proxies for ecosystem health and functioning as they act as indicators of environmental change (Widdicombe and Austen 1998; Peterson and Heck 2001; Lohrer and others 2004; Webb and Eyre 2004; Thrush and others 2006). Development of assessment tools for ecosystem-based monitoring and management is becoming increasingly important as global anthropogenic pressures on natural environments continue to grow and management focus shifts to address issues of loss of function and concomitant ecosystem services. Therefore, understanding organism–sediment interactions and consequently the underpinning mechanisms by which changes across environmental gradients are mediated, is essential for better prediction and interpretation of ecosystem functioning (Williamson and others 1999; Suding and others 2008).

Soft-sediment environments are some of the most widespread habitats on the planet incorporating most of the world’s coastal zones and estuaries (Snelgrove 1999). By physically reworking the sediment they live in, bioturbators increase the energy flow across the sediment–water interface, subducting organic, and labile material whilst introducing oxygen to greater depths than would otherwise be the case (Aller and Yingst 1978; Kristensen and others 1985; Hollertz and Duchêne 2001; Botto and Iribarne 2000). In shallow water ecosystems, pelagic primary production is reliant on benthic processes that recycle nutrients back to the overlying water column (Sundback and others 2003; Gibbs and others 2002). Many of these systems have sufficient light penetration to support benthic primary production by microphytobenthos, tightly coupling photosynthesis with the flux of inorganic nutrients to the water column (Sundback and others 1991, 2000). Bioturbators can positively affect microphytobenthic populations by enhancing mineralization rates and increasing availability of ammoniacal nitrogen, or negatively affect production through grazing and subduction (Lohrer and others 2005; Tang and Kristensen 2007). Therefore, the interactions between benthic primary producers, bioturbation and the sediment characteristics should be accounted for to correctly interpret estuarine and coastal ecosystem processes in sedimentary systems.

Even the most ubiquitous macrobenthic species show some habitat preferences contributing to changes in densities across environmental gradients (Ysebaert and Herman 2002; Thrush and others 2003). The degree to which increased proximity to one another influences ecosystem processes is, in part, governed by an organism’s biological or functional traits (trophic guild, mobility, and lifestyle mode). For example, burrow-building species have been shown to alter the diffusion dynamics of porewater solutes through their proximity to one another (Aller and Aller 1998; Gilbert and others 2003), whilst differences in grazing intensity at the sediment surface can alter the direction and degree of sediment–water solute exchange through the impact on microphytobenthic populations (Marinelli 1992). Increased organism density and therefore increased availability of metabolic products such as NH4 + can also stimulate microbial metabolism, affecting nitrification/denitrification rates (Henriksen and others 1983; Kristensen 1985). Hence, changes in organism density can exert strong control over ecosystem processes such as nutrient cycling (Widdicombe and Austen 1998; Lohrer and others 2004; D’Andrea and DeWitt 2009; Sandwell and others 2009).

Often, species are categorized to a functional group by their biological traits to infer wide scale patterns in ecosystem functioning (Norling and others 2007; Suding and others 2008; Bremner and others 2006). By definition, some functional traits indicate a shift in organism behavior such as the facultative switch between feeding modes in some benthic organisms under differing environmental regimes (Riisgård and Kamermans 2001; Marinelli and Williams 2003). However, few studies to date have considered that single species’ can perform differing ecological functions dependent on their sediment environments, a phenomenon likely to be of particular importance in systems where single species dominate the biomass (Sassa and Watabe 2008). Rarer still are those studies which combine interactions between habitat, organism behavior, and density to infer changes to ecosystem processes (McCraith and others 2003; Escapa and others 2008). Here we consider not only context-specific functioning of organisms but also their density-dependent effects, to create more robust models of organism–sediment relationships and their associated biogeochemical processes.

Burrowing crabs play important roles in ecosystem functioning (Lee 1998; Kostka and others 2002; McCraith and others 2003; Wang and others 2010). They are often found in high densities and can significantly alter the sediment–water interface through burrow construction modifying bedload transport dynamics, affecting sediment penetrability, erodibility, and increasing sediment mixing (Botto and Iribarne 2000; McCraith and others 2003; Gutiérrez and others 2006; Escapa and others 2007; Needham and others 2010; Wang and others 2010). Here we demonstrate how the mud crab Austrohelice crassa (herein Austrohelice), can have context-specific effects on biogeochemical processes through sediment induced changes to its primary bioturbational role. Shifts in functionality are mediated through differences in burrow permanency and consequentially, sediment reworking rates between cohesive and non-cohesive sediments (Needham and others 2010). In situ benthic chambers were used to measure the rate of solute flux (O2, NH4 +, PO4 3−, NO3 −, NO2 −) in both sand (S) and muddy-sand (MS), in relation to Austrohelice density. Flux incubations were conducted in both daylight and at night to establish the relationships between microphytobenthic photosynthetic O2 production, microphytobenthic nutrient uptake and crab bioturbation in the two sediment types. We hypothesized that the observed differences in sediment reworking rates would be detectable in the flux of solutes across the sediment–water interface due to the differing mechanisms facilitating their utilization or release.

Increasing crab density was predicted to reduce and potentially alter the microphytobenthic community through increased grazing pressure, in-turn affecting sediment oxygen concentrations, gross primary production and nutrient release irrespective of sediment type. However, changes in sediment reworking rates and hence Austrohelice’s primary bioturbational role between sand and muddy-sand sediments were predicted to create significant between site differences in the magnitude of solute fluxes at equivalent crab densities. Due to the transient nature of burrows in sand (lasting ~1.4 days, Needham and others 2010) collapse and subduction of microphytobenthos was anticipated to reduce productivity compared to the more physically stable muddy-sand burrows (lasting ~11 days, Needham and others 2010). Regular destabilization of the sediment matrix through burrow collapse may also alter diagenic pathways and processes in sandy sediments, enhancing remineralization rates, and porewater flux in association with increasing Austrohelice density. As burrow longevity is greater in muddy-sand, these structures likely create diverse, well established microbial communities within the burrow walls. Such communities exert strong chemical control over biogeochemical processes altering the pathways and magnitude of solute fluxes. In the case of nitrogen recycling, the close coupling of nitrification/denitrification pathways in burrow walls has been well documented (Kristensen 1985 and refs therein; Gilbert and others 1998; Webb and Eyre 2004). Having said this, uncoupled nitrogen dynamics through burrowing crab activity has also previously been seen (Botto and others 2005). Potential rates of nitrification/denitrification and nitrate reduction generally reflect the state of the microbial population. We therefore predicted that stronger relationships between these processes would occur in muddy-sand sediments compared to those in sand, and that the increased sediment–water interface associated with greater crab (and hence burrow) density, would further amplify reaction rates.

Methodology

Study site and Experimental Treatments

This study was conducted in Tairua estuary (36°59′57.56″S, 175°51′04.09″E), Coromandel Peninsula, New Zealand, in the mid-intertidal zone during summer 2008. The two sediment types, S (sand) and MS (muddy-sand) were chosen as differences in crab behavior affecting sediment stability had previously been recorded in these sites (Needham and others 2010). Site S was categorized as a fine to medium sand and site MS as a fine sand with high silt–clay content (see results). In each sediment type a randomized complete block design was employed, whereby 16 cages (60 × 60 × 40 cm, L × W × H) were installed in blocks of four spanning a total distance of about 45 m parallel to the incoming tide. Crab density treatments of 0, 15, 25, and 35 ind. 0.36 m−2 cage, equating to densities of 0, 42, 70, 98 ind. m−2, were created with one replicate in each of the four blocks (n = 4). We used only mature crabs with an approximate carapace width (CW) of 10 mm, which were collected from the same location within Tairua estuary. Our high treatment density was less than the maximum observed in natural Austrohelice populations of this estuary; up to 252 ind. m−2 in muddy-sand and up to 190 ind. m−2 in sand have previously been recorded (Needham and others 2010). However, these estimates included a high proportion (~50%) of juveniles (CW < 5 mm) that did not create large, deep burrows and therefore had limited bioturbation potential. Compared to the peak densities of larger adult crabs (>8 mm CW; S = 71, MS = 86 ind. m−2; Needham and others 2010, H. Needham pers obs.), our maximum treatment density was slightly higher than the maximum observed density, to buffer against potential crab loss (see results).

To establish each cage, sediment was excavated to 20 cm depth (that is, well below the average burrow depth, Needham and others 2010). The excavated sediment was sieved on a 2 mm mesh screen to remove all large macrofauna. Each cage (constructed from a single piece of 3 × 6 mm nylon mesh (Taylor Built Ltd) and woven at the seams) was set into an excavated hole, leaving the top 20 cm of the cage walls visible above the ambient sediment surface. Sieved sediment was placed back inside each cage and leveled before weaving on an inset lid. As coarser sediments are more adapted to periodic resuspension and disturbance, site S was left for 2 weeks whilst site MS was left for 3 weeks, to re-establish chemical gradients, microphytobenthic, and faunal communities prior to crab introduction (Davis and Lee 1983).

Individual Austrohelice were randomly assigned to cages according to the treatment densities and allowed 18 days to establish themselves at site S and 22 days at MS prior to flux measurements. This allowed the crabs to rework the sediment into a more natural state as sediment at site MS showed greater burrow density due to increased burrow longevity (Needham and others 2010). Cages were monitored regularly for signs of sediment alteration, physical damage, crab mortality, and burrow number. No alterations to the sediment surrounding the cages were observed in either sediment type and burrow number appeared proportional to each treatment in all but one instance (see results).

Flux Incubations

Benthic chambers at S and MS were deployed on consecutive spring tides. At high tides, the depth of water covering the chambers was approximately 1.25 m. Incubations were paired (daylight and night) to account for interactions with microphytobenthos. All water from within the chambers drained with the tide between the two incubation periods. Chamber bases (0.25 m2) were placed within the 0.36 m2 cages 12 h prior to running incubations. Perspex domes were fitted on an incoming tide, entrapping on average 36 l of water. A recirculating water pump intermittently mixed the chamber water to stop stratification without resuspending sediments. Midge™ oxygen loggers (Eureka Environmental Engineering, Texas) were placed in a cradle on a wall of each chamber away from sampling ports to measure water oxygen concentration at 5 min intervals. As biological reactions and solute exchanges can be driven by alteration in temperature and light levels, three Hobo loggers and eight tidbit loggers (Onset computing corporation) logging every 5 min, were randomly assigned to chambers across treatments. Additional temperature and light sensors were placed on the sediment surface at an unmanipulated ambient water sampling station located between the chamber blocks.

Water samples were collected through 2 m of 3.2 mm dia. nylon tube capped with a luer lock valve, to minimize sediment disturbance. A one-way valve, placed on the opposing wall to the sampling port allowed external water to be drawn into the chamber to compensate for sample water removal. Ambient water samples, drawn from the ambient water station close to the chambers, were used to correct for nutrients drawn back into the chamber. Once the chambers were sealed, one clear and one dark 1 l bottle were filled with ambient water and secured at the sediment surface to enable water column processes at the experimental site to be factored out of flux calculations.

Sampling began once all chambers were sealed and fully submerged. After discarding the first 20 ml to clear the tubing, 50 ml water samples were collected via syringe every 45 min until the tide receded 4–5 h later. After each sampling period, water samples were pressure filtered through a swinnex housed 24 mm Whatman GF/C filter into an acid washed container. Samples were kept on ice in the dark prior to freezing. Solute fluxes were determined through linear regression of concentrations as a function of time, chamber volume, and sediment surface area. These rates were corrected for the addition of replacement water drawn into each chamber at the time of sampling.

After diurnal incubations were completed, two burrows within each chamber were randomly selected and cast using a catalyzed polyester resin (Norski products) to check that burrow morphology had not been affected by caging. Sediment reworking rates were also calculated from these burrow volumes. A 13 cm dia., 15 cm deep macrofaunal core (0.01 m2) was also taken from each chamber and sieved over a 500 μm mesh to determine the biomass and identity of any other bioturbators present. Three 28 mm dia., 1 cm deep sediment cores were collected and pooled from each chamber and the ambient station (where Austrohelice were also present) for granulometry, sediment pigment, and total organic matter content (TOM). All samples were frozen in the dark immediately post-collection. The sediment from within each chamber as well as the perimeter material (that is, that still enclosed in the cage but not included within the chamber) was then excavated and sieved separately on a 2 mm mesh to collect the crabs.

Laboratory Analyses

Inorganic nutrient species (NH4 +, NO3 −, NO2 −, and PO4 3−) were analyzed colorimetrically on a Lachat CQ8000 Flow Injection auto analyzer (DKSH Ltd) using standard Lachat QuikChem® methods. TOM was determined through loss on ignition from dried sediments (110°C for 24 h), after combustion for 5.5 h at 550°C (Dean 1974). Sediment grain size fractions were determined using a Malvern mastersizer-S (300 FR lense, range 0.05–2000 μm) after digestion in 10% hydrogen peroxide (Day 1965). Pigments were extracted from freeze-dried sediment and steeped in 90% acetone for 24 h before centrifugation. Chlorophyll a (chl a) and phaeophytin (phaeo) content was determined flourometrically, before and after acidification on a Turner 10-AU flourometer (Arar and Collins 1997). Macrofauna were stained using rose bengal, grouped by major taxa, counted and blotted wet weighed (BWW). All crabs were sized (CW) using digital calipers and counted. Burrow casts were classified and analyzed morphometrically as described in Needham and others (2010).

Data Analyses

Due to the variability in the number of Austrohelice collected from the chambers post-incubation (see results), analysis was conducted using multiple linear regression with backwards selection to determine the effects of crab flux chamber density and sediment properties on dissolved solute fluxes. Light intensity, temperature, and macrofauna BWW (excluding crabs) were also included as predictor variables in initial (full) models. Multicolinearity among predictor variables was avoided by examining the variance inflation factors and condition indices. Variables were eliminated using a backwards selection procedure (SAS 9.1.3) unless significant at α = 0.10. The significance of final models was evaluated at α = 0.05, and goodness of fit was assessed using adjusted r 2 and Aikaike’s Information Criterion (AIC) values. Assumptions of homogeneity of variance and normality were evaluated by plotting residuals versus predicted values, with normal probability plots and Shapiro–Wilk’s tests on residuals. Homogeneity of slopes tests were conducted to elucidate if density-dependent effects were similar between the two sediment types. If no interaction occurred, analysis of co-variance (ANCOVA) was used to determine the main effects (site/density). Any significant co-variables highlighted in the multiple linear regressions were omitted from this analysis and corrected partial residual values were used where necessary. Sediment reworking values were derived using formulas and burrow permanency data from Needham and others (2010). Mann–Whitney non-parametric U tests were conducted to look at differences in reworking, burrow surface area and burrow:crab ratios between sediment types, as the data were not normally distributed.

Results

Sediment Properties and Crab/Sediment Relationships

Sediment properties in the cages were not substantially different from the surrounding ambient sediment at either site indicating there were no long-term effects of the establishment procedure (Table 1). Between site differences in median grain size and silt–clay content were as expected, with many of the other sedimentary variables similar across sites (see Supplementary Material Table A.1 for details). Correlations between Austrohelice density and surface sediment properties showed similar trends at both sites, but the strength of correlations varied (Table 2). At both sites crab density was negatively correlated with sediment pigments (chl a P = 0.001–0.02, phaeo P = 0.03–0.02), TOM (not significantly in S P = 0.11, marginally at MS P = 0.049), and silt–clay content (site S, P = 0.001) though not significantly in MS (P = 0.14) whereas median grain size was positively correlated with crab density at both sites (P = 0.03–0.04).

Table 1 Mean (SD) Sediment Properties of Caged and Ambient Sediments at Each Site
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Table 2 Correlation (Pearson’s Coefficient) Between Austrohelice Density and Surface Sediment Properties
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Burrow structures were not affected by the cage enclosures. All burrow casts were the same morphological forms as previously described from these sites, most commonly ‘i’ an often short, oblique angled straight burrow (Site S = 67%, MS = 39%) and ‘j’ an oblique angled burrow with a terminal hook or chamber (site S = 17%, MS = 30%). As seen previously, burrow morphotypes showed no correlation with either crab density or sediment type (Needham and others 2010). Data across treatments were pooled at each site (n = 24 casts), to assess if burrow surface area and sediment reworking rates altered as a function of sediment type. Burrows had a greater median depth at site S (6.1 cm) than MS (5.0 cm), as well as greater mean volume (S = 48 ± 21 cm3 (±SD), MS = 25 ± 10 cm3) and surface area (91 ± 35 cm2 and 58 ± 18 cm2, respectively). At site S, on average 57 ± 30 kg m−2 (±SD) each summer lunar month (SLM), was reworked opposed to only 3.6 ± 1.7 kg m−2 SLM−1 at site MS. This equates to a 16-fold increase in sediment reworking at site S. Mann–Whitney U tests showed significant differences in burrow surface area (P = 0.009) and sediment reworking rates (P < 0.001) between the two sediment types.

Organism Density and Biomass

Evidence of crab mortality or cage breach was not obvious from the sediment surface. However, from the final number of Austrohelice recovered from each cage, losses (and occasional gains) occurred in both sediment types and across density treatments (Table 3). On average 83% of the total number of crabs recovered from a given cage were found inside the incubation chamber, a value comparable to the proportion of the cage area occupied by the chamber (70%). This indicates that crabs did not avoid or aggregate at the perimeter of the cage and therefore the presence of the cage did not greatly influence burrow position. Data from one flux chamber in MS were omitted from all analyses due to the presence of three thalassinidean shrimp. With no obvious sign of disruption to the cage or lid observed, a breach of the mesh cage walls below the sediment surface most likely occurred. Correlations between burrow and crab density were only seen at site MS (Pearson’s correlation P = 0.019, at site S P = 0.15). Overall burrow:crab ratios were lower in site S than MS (Mann–Whitney U test, P = 0.02) indicating that burrow permanency (and therefore the number of refuges available) was greater at MS, a phenomenon previously recorded in the natural environment (Needham and others 2010). This suggests that crab behavior was not greatly compromised by the presence of our cages. Both biomass (BWW) and abundance of other macrofauna within the caged sediments were dominated by small polychaetes, which were able to migrate through the mesh cage.

Table 3 Initial and Final Austrohelice Densities and Benthic Community Composition
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O2 Fluxes and Primary Productivity Estimates

Leaks, through improper attachment of the chamber lids at night were identified in three of the chambers at site S (crab densities 0, 56, and 116 m−2) and one chamber at site MS (108 crabs m−2). These measurements were therefore excluded from all analyses.

At both sites sediments acted as a source of oxygen during daylight and a sink for oxygen at night (Figure 1). Multiple linear regression models for both daylight and night incubations in each sediment type were significant and explained 27–57% of the variability in O2 flux, with crab density being the greatest significant predictor at both sites (Table 4). During daylight the slope of the relationship between crab density and O2 efflux did not differ with site (Table 5). However, a significant between site difference in O2 flux was observed (P ≤ 0.001; Table 5), with rates approximately 3.8 times greater in site S than MS (adjusted means, S = 1873 and MS = 496 μmol O2 m−2 h−1). A significant crab density × site interaction at night (f = 4.47, P = 0.046) indicated that differences in the relationships between crab density and O2 flux between sediment types occurred. During the night when microphytes were no longer photosynthesizing, O2 influx into both sediment types was observed in all chambers (Figure 1). O2 influx increased with increasing crab density due to greater respiration demands, being most pronounced at site S where TOM was also negatively correlated with O2 flux (Table 4 P = 0.08). However, partial residual plots without this variable showed 61% of the variation was explained by density alone.

Figure 1
figure 1

O2 flux as a function of Austrohelice density (using residual data where appropriate, see Table 4) at site S (square) and MS (triangle). Daylight is denoted by open symbols and night by closed symbols. Regression line is solid for site S (day: y = −2513 − 53x, r 2 = 0.31, P = 0.024, night: y = −254 − 53x, r 2 = 0.61, P = 0.011) and dotted for site MS (day y = −44.4x + 954, r 2 = 0.30, P = 0.026, night y = −594 − 44.4x, r 2 = 0.27, P = 0.041). Positive values (efflux) indicate sediment release and negative values (influx) imply sediment utilization.

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Table 4 Significant Multiple Linear Regression Models of the Effects Crab Density and Sediment Properties on Dissolved Solute Fluxes
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Table 5 Comparisons of Solute Fluxes Between Sites in Daylight and at Night
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Austrohelice density was negatively correlated with chl a concentration in both sediment types (Table 2). The slope of the regression was greater in site MS than S (0.25 vs. 0.16) but the difference was not significant (homogeneity of slopes P = 0.34). When gross primary production estimates (GPP = daylight O2 flux − night O2 flux) were normalized for microphytobenthic biomass (that is, sediment chl a content), crab density was shown to have a positive, marginally significant (P = 0.05) effect on GPP (Table 5; Figure 2). Between site differences were more significant (P = 0.01) with adjusted means of 292 and 204 μmol O2 μg g−1 dw chl a m−2 h−1 in site S and MS, respectively.

Figure 2
figure 2

Gross primary production (GPP, after normalizing data for microphytobenthic biomass) as a function of Austrohelice density at site S (square) and MS (triangle). Regression line is solid for site S (y = 274 + 2.1x, r 2 = 0.31, P = 0.056) and dotted for site MS (y = 132 + 7.7x, r 2 = 0.34, P = 0.045).

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Nutrient Fluxes

Austrohelice density was the only significant predictor of NH4 + flux, explaining between 52 and 67% of the variance in the regression models (Table 4). Both sediments showed greater NH4 + fluxes with increasing crab density in daylight and at night. However, homogeneity of slope tests showed significant density × site interactions, demonstrating that between site differences in the effect of Austrohelice density on NH4 + flux were detected (Table 5). Between site differences were more significant during daylight (f = 17.5, P ≤ 0.001), than at night (f = 5.5, P = 0.028) with fluxes being over 4.3 times greater at site MS than S in daylight at the highest crab densities. In daylight, NH4 + was taken up by both sediments until Austrohelice density exceeded approximately 60 crabs m−2 (Figure 3A). After this point (in most instances), NH4 + was released to the water column to a greater (MS) or lesser (S) degree. At night, sediment influx of NH4 + also occurred in crab exclusion plots in both sites, despite the absence of photosynthesis.

Figure 3
figure 3

Nutrient fluxes as a function of Austrohelice density at site S (square) and MS (triangle). Daylight is denoted by open symbols and night by closed symbols. Regression line is solid for site S and dotted for site MS. Positive values (efflux) indicate sediment release and negative values imply sediment utilization (influx). A NH4 + flux (S day: y = −33.1 + 2.93x, r 2 = 0.67, P ≤ 0.001, night: y = −27.62 + 4.84x, r 2 = 0.65, P ≤ 0.001. MS day: y = −144 + 14.6x, r 2 = 0.60, P ≤ 0.001, night: y = −37.6 + 18.8x, r 2 = 0.52, P = 0.003). B NO3 − flux; no significant relationships with Austrohelice density were found. C PO4 3− flux; only site S (night) showed a significant relationship with Austrohelice density (y = −10.63 + 0.83x, r 2 = 0.34, P = 0.021).

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NO3 − and NO2 − flux did not show any significant diurnal patterns associated with crab density or sediment type (Table 4). Other significant models for predicting patterns in NO2 − and NO3 − fluxes were influenced by sediment properties such as TOM, silt–clay content and porosity. At site MS NO3 − was nearly always taken up by sediment (Figure 3B), displaying a greater influx in daylight (mean = −46.9 ± 35.4 SD) than at night (−5.02 ± 6.5), yet fluxes of NO3 − and NO2 − (in combination, NO x ) in both sediments were very low. Similarly , fluxes of PO4 3− were low compared to that of NH4 + and O2 (Figure 3C). Crab density did not significantly affect PO4 3− fluxes during the day in either sediment type and variability was not further explained by any of the measured parameters (Table 4). At night, PO4 3− (log10 PO4 3−) efflux was positively influenced by crab density in site S only, increasing with sediment oxygen demand.

Discussion

The presence of Austrohelice was linked to differences in solute fluxes between sand (S) and muddy-sand (MS) environments, altering ecosystem functioning between habitats. Significant density-dependent effects on O2 and NH4 + flux were also apparent. However, variation in the magnitude of change associated with crab density between sites highlighted the different interactions between Austrohelice and its environment in each sediment type. These changes related to the functional attributes of Austrohelice that were modified by environmental conditions. Patterns in both burrow:crab ratios and sediment reworking rates across treatments showed significant between site differences. Greater silt–clay content and hence increased sediment cohesion in muddy-sand (Table 1) resulted in longer lasting burrow structures and a greater burrow: crab ratio overall. Although burrow dimensions and therefore sediment–water interface were greater in sand, their transient nature is likely to weaken their influence on sediment geochemistry because if burrows are short-term, microbial communities cannot easily establish (Marinelli and others 2002). Such differences highlight the context-specific nature of Austrohelice bioturbation on ecosystem processes.

At both sites, crab density influenced sediment properties in similar ways, but to differing degrees (Table 2). The sand site displayed a highly significant negative correlation between silt–clay content and crab density consistent with the shift in functionality between sites. Increased sediment reworking is often associated with a winnowing of fine particles which are easily transported with the tide (Graf and Rosenberg 1997). However, such clear correlation in muddy-sand may have been masked by greater variability in the (naturally higher) silt–clay content. As grazing pressures were assumed to be the same across both sites at equivalent Austrohelice densities, the more significant negative correlation between chl a and crab density in sand is likely to be influenced by increased microphyte reburial through burrow collapse and construction. Nevertheless, GPP was enhanced with increasing crab density when differences in microphytobenthic biomass, through increased deposit feeding were accounted for (Figure 2). This indicates that microphytes were more productive at greater crab densities, likely driven by the increased availability of NH4 +, the preferentially used nitrogen source by microphytobenthos, coupled with a potential release from CO2 limitation through increased crab respiration. Also a reduction in microphytobenthic biomass may decrease intraspecific competition for resources between microphytes, increasing productivity (Morrisey 1988). As chl a content was used as a proxy for microphyte biomass it is also possible that increased grazing pressure may have changed the microphytobenthic community composition between sites, altering the productivity per pigment concentration through species shifts (Falkowski and Kiefer 1985). Top-down grazer control on microphyte biomass, resulting in lower benthic primary production, has been well documented in estuarine systems (Asmus and Asmus 1985; Andersen and Kristensen 1988; Webb and Eyre 2004), yet studies which estimate microphyte efficiency in the presence of macrofauna provide greater insight into benthic interactions. Indeed, previous studies which consider this have also concluded that benthic primary production can be stimulated though macrofaunal bioturbation, particularly in the case of vertical mixers (Lohrer and others 2004; Sandwell and others 2009; Rodil and others 2011).

Between site differences in (normalized) GPP were driven by the greater efflux of O2 during the day in the sand site (Figure 1). Differences in light attenuation between sediment types through variations in physical properties such as grain size will influence microphytobenthic photosynthesis and diffusion of O2 from the sediments. Light penetration and backscatter is greater in coarser sediment than fine, enabling photosynthesis to occur at greater depth in the sediment profile (Paterson and others 1998). Bioturbation has previously been seen to have a greater influence on oxygen consumption in diffusion-dominated muddy sediment opposed to advective-dominated sands. Mermillod-Blondin and Rosenberg (2006) demonstrated that U-shaped burrow builders increased O2 consumption in diffusive systems due to their strong influences on microbial processes. The opposite effect was true in more permeable sediments where changes in water circulation only moderately influenced microbial processes. Microphytobenthic species assemblages are also likely to differ with grainsize, altering community production rates between locations (Heip and others 1995). However, relationships between Austrohelice density and O2 flux in daylight were similar in both sites despite fluxes being nearly four times greater in sand than muddy-sand, indicating crab respiration dominated sediment O2 demands (Figure 1).

At night O2 influx increased with crab density in both sediments due to increased respiration pressures within the chambers (Figure 1). Austrohelice does not show strong circadian rhythms, but instead shows peak activity around high tide (Williams and others 1985). Therefore, organism behavior was assumed to be similar in both day and night chamber deployments. Nonetheless, differing interactions between sediments and crab density were observed between sites at night, with the greatest influx of O2 occurring in site S. TOM was a significant predictor of O2 flux in site S which, in conjunction with the increased availability of NH4 + and greater porewater exchange, indicates some enhancement of remineralization through bioturbation activity. Vertical mixing of sediments has been well documented for enhancing remineralization rates through the constant movement of particles between reaction zones, increasing overall benthic metabolism (Aller 1994). Increased remineralization and sediment mixing through bioturbation can also influence the adsorption/desorption pathways of PO4 3− (Slomp and others 1998). However, the increased PO4 3− efflux in sand at night (Figure 3C) may also be due to the release of porewater nutrients associated with higher crab densities and hence, sediment reworking. This effect may be masked in daylight due to microphyte utilization.

Benthic organisms excrete ammonium through metabolic activities, contributing to the overall flux of NH4 +. Increased NH4 + flux with increasing crab density was observed in both sediment types during daylight and at night (Figure 3A). Similar flux rates between sites in crab exclusion plots in both daylight and at night indicated that sedimentary driven differences were minor. Where crabs were excluded or present in low numbers, NH4 + was taken up by the sediment, indicating utilization by microphytes and bacteria. However, with increasing crab density, excretion rates are likely to have outweighed autotrophic demand leading to a release of excess NH4 + to the overlying water. This shift from influx to efflux occurs at densities above 60 ind. m−2 in both sediment types and is more pronounced in muddy-sand where O2 efflux and GPP were generally lower (Figures 1, 2). Despite similarities in the fluxes between sediments when crabs were excluded, NH4 + fluxes were of greater magnitude in muddy-sand than sand diurnally, indicating that these changes in responses were Austrohelice mediated.

Crab nitrogen excretion rates were assumed to be the same at each site. Consequently, the amplified efflux of NH4 + at site MS was attributed to stimulation of NH4 + production through microbial mineralization pathways within the crab burrow walls (Braeckman and others 2010). Established burrows provide a stable interface for these geochemical exchange gradients to develop (Aller and Yingst 1978), unlike the more transient burrow structures in sand. In cohesive sediments, where diffusive processes dominate, burrows act as macropores that produce strong fluxes at the sediment–water interface (Mermillod-Blondin and Rosenberg 2006). Flows over these reactive burrow surfaces greatly influence microbial processes. With increased crab density, galleries of burrows were formed in site MS amplifying these reactions and further enhancing N-cycling. Irrigation through persistent crab movement into and out of burrows, as well as passive irrigation through tidal flow in the natural system, may also affect NH4 + efflux to some degree by creating pressure differentials (Williams and others 1985; Ray and Aller 1985).

Similar NO3 − influxes between sites were also observed where crabs were excluded, the magnitude of which were equivalent in daylight and at night, yet Austrohelice density driven trends were not detected at either site (Figure 3B). Persistent, albeit low levels of NO3 − influx in muddy-sand, indicated that dissimilatory nitrate reduction may have been promoted in anoxic areas of burrow walls (Kristensen and others 1985), even when O2 production was enhanced at the sediment surface. However, burrow morphology was not correlated with Austrohelice density in either site. Uptake by microphytes may also have contributed to NO3 − influx in muddy-sand during daylight, particularly in lower crab densities where NH4 + influx was also measured. Low NO3 − and NO2 − (NO x ) fluxes at these sites were not unexpected as denitrification rates are often witnessed to be greater than nitrification rates in marine sediments, resulting in little available NO3 − to diffuse out of sediments. Nitrogen efflux in coastal habitats therefore consists mainly of NH4 + (Kemp and others 1990; Trimmer and others 1998; Thrush and others 2006) and in New Zealand, NH4 + has previously been seen to make up 100% of the measured dissolved inorganic nitrogen in coastal systems (Lohrer and others 2004, 2010; Sandwell and others 2009).

Austrohelice is a key bioturbating species in the regulation of nutrient cycling and remineralization rates in both muddy-sand and sandy environments and shows strong density driven changes to solute exchange. Changes in the bioturbational role of Austrohelice were most notable in the flux of NH4 +; the key source of inorganic nitrogen in these systems. Austrohelice bioturbation also facilitated higher benthic primary production per unit of chl a in both sediment types, increasing with crab density. This unpredicted result may be facilitated through an increase in Austrohelice excretory products (NH4 + and CO2) with greater crab density, which may act as a trade-off against increased grazing pressure.

The importance of faunal activities for the regulation of ecosystem processes is frequently associated with individual species and their functional or biological traits (Bonsdorff and Pearson 1999; Norling and others 2007; Bremner and others 2006). Other studies have concluded that differing bioturbation modes do not have comparable effects on biogeochemical processes under different sediment conditions due to hydrological shifts between environments (Mermillod-Blondin and Rosenberg 2006; Volkenborn and others 2010). Our study highlights that an organism’s contribution to ecosystem functioning can also be context-specific and although different sediment properties can create changes in fluxes, this is an oversimplification of the processes at work. How widespread the phenomena of habitat induced functional plasticity actually is for other species, requires further exploration. Integrating behavioral information into functional studies will broaden conceptual frameworks and further develop accurate predictive tools for ecosystem-based monitoring and management.