Response of temperate intertidal benthic assemblages to mangrove detrital inputs

Journal of Experimental Marine Biology and Ecology 460 (2014) 80–88

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Response of temperate intertidal benthic assemblages to mangrove detrital inputs Rebecca V. Gladstone-Gallagher a,⁎, Carolyn J. Lundquist b,c, Conrad A. Pilditch a a b c

School of Science, University of Waikato, Private Bag 3105, Hamilton, New Zealand National Institute of Water and Atmospheric Research Ltd. (NIWA), PO Box 11115, Hamilton, New Zealand Institute of Marine Science, University of Auckland, PO Box 349, Warkworth, New Zealand

a r t i c l e

i n f o

Article history: Received 4 December 2013 Received in revised form 4 June 2014 Accepted 9 June 2014 Available online xxxx Keywords: Avicennia marina subsp. australasica Benthic macrofauna Decomposition Detritus New Zealand Spatial subsidy

a b s t r a c t Inputs of macrophyte detritus to soft-sediment habitats can be an important energy source regulating benthic community structure. In the tropics, mangrove detritus forms an essential food source for benthic invertebrates; however, it is unknown whether the same dependence occurs in temperate systems where mangrove detrital inputs to estuaries can be considerable. We investigated whether mangrove detrital deposition to temperate intertidal flats represents a cross-boundary subsidy of organic matter by structuring benthic macro-invertebrate communities on adjacent intertidal flats. To determine whether community responses to detrital deposition were spatially consistent, mangrove detritus was added (260 g m−2, equivalent to summer litter production) to two intertidal sites (with differing background sediment properties and macrofaunal community structure). Subsequent changes to the benthic macrofaunal community and sediment properties were monitored for 6 months following the addition. Benthic community responses to the detrital enrichment were similar at both sites; responses were subtle and involved only small changes in the relative abundances of a few dominant taxa (primarily a reduction in the numerically dominant spionids), rather than major shifts in community composition. The subtle response to such a relatively large detrital input suggests that mangrove detritus in temperate estuaries plays a minor role in shaping the communities on intertidal flats. We suggest that the slow decay (low bioavailability) and relatively low productivity of temperate mangroves result in communities that are less reliant on mangrove detritus, compared to those in the tropics where rapidly decaying mangrove detritus comprises the base of many food webs. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Macrophyte detritus (dead, decaying organic matter) is a key source of energy input to many estuarine systems (Findlay and Tenore, 1982; Moore et al., 2004; Odum and Heald, 1975). Deposited detritus can be utilised directly by benthic detritivores (Findlay and Tenore, 1982; McClelland and Valiela, 1998; Moore et al., 2004), as well as fuel the growth of sediment microorganisms (e.g. Bishop and Kelaher, 2007; Levinton, 1985; Rublee, 1982). In addition, numerous studies in temperate estuaries have demonstrated that benthic macrofaunal community structure is modified by macrophyte detrital enrichment (see Bishop and Kelaher, 2008; Bishop et al., 2010; Kelaher and Levinton, 2003; O'Brien et al., 2010; Rossi and Underwood, 2002; Taylor et al., 2010). Accordingly, detrital deposition and distribution can be a key factor regulating small-scale variability and patchiness in soft-sediment community structure and function (Kelaher and Levinton, 2003; Kelaher et al., 2003, 2013; Rossi and Underwood, 2002).

⁎ Corresponding author. Tel.: +64 7 8384466. E-mail address: [email protected] (R.V. Gladstone-Gallagher).

http://dx.doi.org/10.1016/j.jembe.2014.06.006 0022-0981/© 2014 Elsevier B.V. All rights reserved.

Macrofaunal responses to detrital addition should vary with resident community structure due to species-specific responses. It is well established that macrofaunal community structure varies with sediment properties (e.g. grain size and organic content; Pratt et al., 2013; Thrush et al., 2005; van der Wal et al., 2008), and since these properties can also influence detrital decay rates (Holmer and Olsen, 2002) community response to detrital addition can be expected to vary across sedimentary gradients. However, few field-based studies have investigated the site-specific impacts of detrital deposition on estuarine benthic community structure (these are: Bishop and Kelaher, 2013b; O'Brien et al., 2010; Olabarria et al., 2010; Rossi, 2006; Rossi and Underwood, 2002), and just two of these studies have incorporated differences in sediment properties and associated differences in benthic communities among sites (O'Brien et al., 2010; Rossi and Underwood, 2002). Both of these studies explored community responses to the burial of whole algal wrack, and found that different species responded at mud compared to sand sites (Rossi and Underwood, 2002), and between sites with different organic enrichment levels (O'Brien et al., 2010). Mangrove detritus has been shown to be an important source of energy in tropical coastal ecosystems (Doi et al., 2009; Granek et al., 2009; Odum and Heald, 1975). In temperate New Zealand estuaries, recent

R.V. Gladstone-Gallagher et al. / Journal of Experimental Marine Biology and Ecology 460 (2014) 80–88

changes to catchment land use have altered mangrove distributions, which are likely to continue because of climate change and mangrove management strategies (e.g. forest clearances) (Morrisey et al., 2010). As a result, the magnitude of mangrove detrital inputs into temperate coastal systems is changing. Many tropical coastal systems rely on mangrove detritus as a subsidy of organic matter that supports the base of marine food webs (e.g. coral reefs and estuaries, Granek et al., 2009; Sheaves and Molony, 2000; Werry and Lee, 2005); however, the lack of ecological knowledge gathered in temperate mangrove systems offers little guidance to the impacts of changing detrital inputs on recipient coastal ecosystems. Whilst temperate mangroves can be substantially less productive than their tropical counterparts, detrital inputs are comparable to seagrass production in some estuaries (Gladstone-Gallagher et al., 2014), a detrital source known to be important for benthic invertebrates (Doi et al., 2009). Mangrove leaf litter decay is slow in temperate regions compared to other detrital sources (Enríquez et al., 1993) and involves a two part process, where initial decay is rapid, followed by the gradual decay of the recalcitrant portion of the leaf. Initial decay is likely through bacterial colonisation and breakdown of the leaf, which results in nitrogen enrichment (Gladstone-Gallagher et al., 2014) and increasing palatability to organisms (Nordhaus et al., 2011). The slow, secondary decay is most likely through physical fragmentation of the recalcitrant fractions of the leaf, which is controlled by climatic variables, tidal submergence, and the presence of fauna (e.g. Dick and Osunkoya, 2000; Oñate-Pacalioga, 2005; Proffitt and Devlin, 2005). Differences between physical, chemical and biological properties of different sediment types could therefore influence detrital breakdown rates associated with both stages of decomposition. Previous studies have linked sediment properties with differences in the decay rates of both macroalgae and mangrove litter (Hansen and Kristensen, 1998; Holmer and Olsen, 2002; but see Gladstone-Gallagher et al., 2014). Intertidal soft-sediment communities are dynamic (e.g. Morrisey et al., 1992; Thrush, 1991; Thrush et al., 1994) and a temporally variable response to detrital addition could be expected as the decay process proceeds. However, only a few studies investigating macrofaunal responses to detrital deposition have incorporated temporal sampling into study designs, with most monitoring responses on only one or two sampling dates after addition (often sampling two or more months after the addition; e.g. Bishop et al., 2007, 2010; Taylor et al., 2010). The two studies that have incorporated temporal sampling have demonstrated a strong time dependent response: in one, macrofaunal abundance took 24 weeks to respond to the addition of seagrass detritus (Bishop and Kelaher, 2007); whilst in the other, annelids responded to an Ulva detrital addition after only four weeks (Kelaher and Levinton, 2003). These examples illustrate that macrofaunal responses may be variable through time due to differences in detrital types, quantities, decay rates, and the ambient community composition. Accordingly, studies that are restricted to a single sample date may miss some or all of the community response to detrital addition. Here, we investigate the role of mangrove detrital inputs in structuring intertidal benthic communities in a temperate setting. Mangrove detritus was added at two adjacent sites with different background sedimentary properties and macrofauna. The benthic community response was monitored several times over a six month period. We anticipated that changes in macrofaunal community structure would vary with site and time, because detrital processing/decay would be influenced by site-specific sediment biogeochemistry and species responses to enrichment. 2. Material and methods 2.1. Study site The study was carried out in the northern region of Whangamata Harbour (North Island, New Zealand). The New Zealand endemic

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mangrove Avicennia marina subsp. australasica inhabits 101 ha of the harbour (approximately 22% of the 467 ha harbour area), which has expanded from 31 ha of mangrove forest prior to catchment urbanisation, deforestation and agriculture since the 1940s (Singleton, 2007). Such changes in catchment land use have increased the delivery of terrestrial sediments via streams and rivers into many New Zealand estuaries and mangroves have expanded in response to sedimentation (Morrisey et al., 2010). Two unvegetated mid-intertidal sites were selected: site 1 (37°10′43.0″S, 175°51′36.9″E) is characterised by fine organic-poor sands and the adjacent site 2 (37°10′38.6″S, 175°51′36.5″E) has higher mud content and relatively organic-rich sediments (see Results). The two sites are located 20–40 m down-shore of the mangrove forest edge, occupying similar tidal elevations (0.05–0.25 m above mean sea level) and are separated by an along-shore distance of approximately 50 m. The spring–neap tidal range is 1.71–1.27 m (Hume et al., 2007), and inundation periods at the sites were similar (site 1 = 5–6 h and site 2 = 6–7 h). 2.2. Experimental protocol In early February 2011 (late austral summer), 18 evenly dispersed plots (1.15 m dia., 1.04 m2) were established at each site within a 32 m × 14 m area. Five metres separated each plot. In each of three rows of six plots, we randomly assigned two replicates of the following treatments: detrital addition (DA), procedural control (PC), and control (C) (n = 6 for each treatment). DA plots were enriched with mangrove detritus by finger churning 270 g of detritus (260 g m−2) into the top 3 cm of sediment (as in Bishop and Kelaher, 2008; Bishop et al., 2010; Kelaher and Levinton, 2003). The addition equates to the total amount of leaf litter produced during the productive summer months (November–February) in New Zealand forests, with the timing of the addition coinciding with the end of this production peak (Gladstone-Gallagher et al., 2014; May, 1999; Oñate-Pacalioga, 2005; Woodroffe, 1982). PC plots were finger churned, identical to DA plots, but no detritus was added, and were included in the experimental design to delineate benthic community effects of the one off sediment mixing disturbance from detrital enrichment effects. C plots were left untouched. The mangrove detritus used in the manipulation was prepared by firstly collecting yellow senescent (ready to abscise) mangrove leaves from trees in Whangamata Harbour (January 2011). To simulate natural incorporation of mangrove detritus into the sediments, the leaves were oven dried at 60 °C to constant weight and ground into 2 mm pieces. This drying of leaf material is thought to be comparable to the drying out that a fallen leaf would experience if it fell on a mid-afternoon summer low tide, and was necessary to standardise the amount of detritus added to each plot (Bishop and Kelaher, 2008). Experimental plots were repeatedly sampled for macrofauna (1 × 13 cm dia., 15 cm depth core per plot) and surface sediment properties (photosynthetic pigment content, organic content and grain size) at 2, 4, 8, 12 and 24 weeks following the detrital addition. This monitoring period incorporated a series of sampling dates to determine temporal variability in macrofaunal responses to detrital inputs. The sampling period encompasses that of other studies (e.g. Bishop and Kelaher, 2007, 2008; Kelaher and Levinton, 2003), and is also longer than the half-life of mangrove leaf litter in New Zealand (63–88 days; Gladstone-Gallagher et al., 2014). Sediment samples (3 pooled syringe core samples; 3 cm dia., 2 cm depth) were taken within a few centimetres of each macrofaunal core. To minimise the effect of repeated sampling on the benthic community, macrofaunal cores were taken from different positions within the plots on each sampling date and the resulting core holes in-filled with defaunated sand (Lohrer et al., 2010). Additionally, both sites were sampled for macrofauna and sediment properties on day 0 at 6 randomly chosen locations outside of the experimental plots but within the study area. Macrofaunal cores were sieved over a 500 μm mesh

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sieve and preserved in 70% Isopropyl alcohol (IPA). Sediment core samples were kept in dark, cold conditions immediately after collection, and then stored frozen awaiting later analysis.

which we analysed via t-tests) conformed to assumptions of homogeneity of variances and normality, therefore no transformations were necessary. Univariate analyses were conducted using the STATISTICA software package (Statsoft Inc.).

2.3. Sample analysis 3. Results Sediment samples were homogenised and subsamples taken for analysis of sediment properties. Sediment chlorophyll a (chl a) and phaeophytin (phaeo) concentrations were determined on a Turner 10-AU fluorometer, after pigment extraction (90% buffered acetone) and subsequent acidification (0.1 N HCl) (Arar and Collins, 1997). Sediment grain size (GS) was measured using a Malvern Mastersizer 2000 instrument (particle size range: 0.05–2000 μm), after sediment digestion with hydrogen peroxide (10%) (Singer et al., 1988). Sediment for organic matter (OM) content analysis was dried at 60 °C to constant weight and then combusted at 550 °C for 4 h (OM determined by percentage weight loss on ignition). Macrofaunal samples were stained with Rose Bengal solution and fauna separated then identified to the lowest possible taxonomic level under a stereo microscope. Sorted macrofaunal core samples (i.e. fauna removed) were elutriated using a sugar solution to separate light floating material from heavier shell hash and sediment (similar to methods used in Anderson, 1959). Elutriated material was retained on a 500 μm sieve and assumed to be comprised mostly of detrital material, as fauna had previously been removed. This process was used for DA core samples throughout the experiment, and day 0 samples and C cores from 24 weeks were used to measure ambient detrital material. 2.4. Data analysis In preliminary statistical analyses, the day 0 (initial) samples revealed strong between site differences in sediment properties and macrofaunal community structure. These initial, site-specific differences in community structure persisted throughout the experiment and resulted in significant site × treatment and site × time interactions in subsequent uni- and multi-variate analyses (described below). Recognising these initial differences, we chose to present separate analyses for each site to better focus on the effects of treatment and time, and to better illustrate differences in the response between sites. Non-metric multi-dimensional scaling (nMDS) analysis, using Bray–Curtis similarity matrices, was used on raw macrofaunal abundance data to plot and compare benthic community structure among treatments, as well as through time. We analysed these changes statistically using a repeated measures permutational multivariate analysis of variance (PERMANOVA, Bray–Curtis similarity). Time and treatment were treated as fixed factors (with 5 and 3 levels, respectively), with plot (6 levels) treated as a random factor nested within treatment (Anderson et al., 2008). PERMANOVA pair-wise tests were used to determine where significant treatment and time effects occurred, and SIMPER analysis determined the taxa that contributed to the dissimilarity/similarity in community structure between treatments. The percent dissimilarity to standard deviation ratio (Diss/SD) was also used to determine whether the taxa identified in SIMPER analyses were good discriminating taxa (Diss/ SD N 1.3) (Clarke and Warwick 1994). Raw data were used for multivariate analyses as transformations did not alter the results. All multivariate analyses were performed using the PRIMER (with the PERMANOVA A + addition) statistical software program (Plymouth Routines In Multivariate Ecological Research; Anderson et al., 2008; Clarke and Gorley, 2006). A repeated measures analysis (two-way, fixed factor, repeated measures analyses of variance (ANOVA)) was also used to test treatment effects on univariate variables (sediment properties, and macrofauna taxonomic richness and abundance) through time. Newman–Keuls post-hoc tests were used to determine differences between treatments for each sample date. Raw data (including the initial day 0 samples,

3.1. Sediment properties Day 0 sediment properties were significantly different between the two sites (t-tests, p b 0.05), and these inter-site differences persisted throughout the experiment (Table 1). Sediments at site 2 were characterised by a lower median grain size (131 vs. 198 μm), but a mud (b 63 μm), pigment (chl a and phaeo) and OM content approximately 2 times higher than at site 1. At both sites, sediment properties were unaffected by the addition of detritus (repeated measures ANOVA, treatment effect p N 0.05), but properties did vary temporally (time effect p b 0.05). Although statistically significant, the sediment properties at each site did not vary substantially during the experiment, as indicated by the narrow range of treatment means (Table 1). Even though sediment properties did not differ with treatment, we did detect a difference in the amount of detritus (N 500 μm) between DA and control plots at both sites, which persisted for the duration of the experiment (Fig. 1). Two weeks after the addition, the detritus in DA plots was 8–9 times higher than that in ambient sediment measured on day 0, and by the end of the experiment (week 24) there was still marginally significant higher levels of detritus in DA than in control sediments at both sites 1 and 2 (t-test, p = 0.02 and 0.1, respectively). The decline in detritus from DA plots as the experiment progressed through time was probably associated with losses due to decay and/or tidal transport (Fig. 1). 3.2. Macrofaunal community response At the start of the experiment (day 0) macrofaunal communities differed between sites (PERMANOVA, df = 1, pseudo-F = 22.78, p = 0.001). These differences were driven by variations in the total and the relative abundance of several species, rather than differences in the species that were present. Site 1 had approximately half the macrofaunal abundance of site 2 (101 vs. 175 ind. core−1; t-test p = 0.005), although species richness (~ 14–15 taxa core−1; p = 0.6) was similar at both sites. The average dissimilarity between sites was 58%; three polychaete species were good discriminating species (Diss/SD N 1.3) and accounted for 73% of this difference (SIMPER analysis; Table 2). The spionid Prionospio aucklandica and the capitellid Heteromastus filiformis were more abundant (5–10 × higher) at site 2 than at site 1 and accounted for 47% and 12% of the dissimilarity, respectively. At site 1, the spionid Aonides trifida was much more abundant and contributed 14% to the dissimilarity, with the remaining species accounting for b3.5% of the dissimilarity (Table 2). Community responses at both sites showed highly significant main factor effects of treatment and time in PERMANOVA analyses, but no significant time × treatment interaction (Table 3A and B). Post-hoc pairwise comparisons showed that at site 1, differences between DA plots and both control plots (PC and C) were highly significant (p b 0.002), whilst differences between PC and C communities were not significant (Table 4). Such a result suggests that effects on benthic community responses were associated with the addition of detritus and the procedure of mixing the sediment had no significant effect at site 1. At site 2, however there were differences between all three treatments (p b 0.04), suggesting an effect of both the detrital addition, as well as the procedural disturbance (Table 4). Post-hoc pairwise testing for the effects of time showed that communities were temporally variable at both sites (Table 5), and the lack of a time × treatment interaction indicates that this temporal variability was independent of treatment (Table 3).

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Table 1 Initial (day 0) sediment properties and temporally averaged values as a function of treatment (C = control, PC = procedural control, DA = detrital addition) at (A) site 1 and (B) site 2. In brackets, the range of values observed on four sample dates (between 2 and 12 weeks) is given (24 week sediment samples were lost and therefore are not included). Chl a (μg g−1)

OM (%)

Phaeo (μg g−1)

Median GS (μm)

Mud content (%)

(A) Site 1 Initial C PC DA

1.90 2.18 2.18 2.28

(2.09–2.23) (2.02–2.49) (2.12–2.44)

10.50 9.34 9.64 8.34

(8.33–10.76) (8.51–11.79) (7.14–10.55)

2.79 2.80 3.04 3.30

(2.55–3.00) (2.50–4.11) (2.80–4.18)

197.6 227.7 225.6 231.2

(194.5–252.5) (197.8–241.7) (201.9–245.2)

14.4 12.4 12.5 11.2

(9.2–17.7) (9.5–15.6) (9.1–13.4)

(B) Site 2 Initial C PC DA

4.05 4.20 4.06 4.39

(4.06–4.35) (3.90–4.24) (4.14–4.63)

23.80 19.90 19.72 18.74

(18.16–20.90) (18.15–21.89) (16.28–21.06)

4.52 4.90 5.54 5.78

(4.08–5.62) (4.65–6.01) (4.83–6.52)

130.8 162.0 149.5 155.9

(123.2–183.5) (116.7–180.0) (128.6–184.2)

29.9 25.1 27.4 27.1

(21.7–33.1) (21.5–33.4) (21.4–31.2)

OM = total organic matter content of sediment; Chl a = sediment chlorophyll a pigment content; Phaeo = sediment phaeophytin pigment content; GS = grain size; Mud = % silt/clay (particles b 63 μm).

nMDS ordinations are consistent with and depict PERMANOVA results (Fig. 2). At site 1, PC and C communities cluster together, whilst DA communities are more variable and spread out away from both controls (Fig. 2A). However, this pattern is not the same for site 2, where there is some separation between the three treatments (Fig. 2C). Consistent with PERMANOVA results, the nMDS ordinations for both sites show that communities are temporally variable regardless of treatment (Fig. 2B and D). Univariate measures of community structure provide some insight into the PERMANOVA results. Total abundance and species richness varied significantly with treatment and time (repeated measures ANOVA, p b 0.05; Table A.1) at both sites. However, no time–treatment interaction effect was detected in these tests, which suggests that the temporal variation in abundance and richness was similar between treatments. The number of taxa at site 1 was unaffected by the addition of mangrove detritus throughout the experiment (Fig. 3A), however total abundance did show some response to the detrital addition. At site 1, the DA treatment had significantly lower total macrofaunal abundances than C and PC treatments (Fig. 3B; Newman–Keuls, p b 0.05). The mean number of taxa at site 2 was significantly lower in DA and PC plots compared to C plots, suggesting an effect of the procedural disturbance on species richness (Fig. 3A; Newman–Keuls, p b 0.05). Statistical differences in total abundance at site 2 were different to those observed at site 1. Abundance was reduced in DA and PC plots compared to C plots again indicating a potential procedural (disturbance) effect (Fig. 3B; Newman–Keuls, p b 0.05). The trends in total abundance correspond to multivariate PERMANOVA results, suggesting that

differences in community structure were being driven by changes in abundance rather than species composition. SIMPER analysis revealed that the dominant polychaetes P. aucklandica and A. trifida were good discriminating species (Diss/ SD N 1.3) and contributed the greatest percentages (N 40% cumulative) to the dissimilarities between DA and control treatments at site 1 (Table 6). At site 2, dominant polychaetes P. aucklandica and H. filiformis were the discriminating species (Diss/SD N 1.3) that were responsible for the greatest proportion (N 40% cumulative) of the dissimilarity seen between treatments (Table 6). The significant differences between treatments at both sites were driven mainly by the numerically dominant polychaete species P. aucklandica and it is likely that this species was responsible for the decreases in overall abundance (compare Figs. 3B with 4). Decreases in P. aucklandica were associated with the addition of detritus at site 1, as well as the disturbance of mixing the sediment at site 2. Hence, P. aucklandica decreased significantly in DA plots compared with controls at site 1 (Fig. 4). However, at site 2 P. aucklandica decreased significantly in DA and PC plots (Fig. 4). 4. Discussion We added mangrove detritus to the sediment on two unvegetated intertidal flats, where it remained throughout the 24 week experiment. Several studies have manipulated detrital additions to determine the role of different detrital sources in creating small-scale variability in soft-sediment benthic communities (e.g. Bishop and Kelaher, 2008; Bishop et al., 2010; Kelaher and Levinton, 2003). However, the use of

Fig. 1. Mean dry weight of detritus (N500 μm; ±1 SE, n = 6) remaining in detrital addition plots (DA; filled circles) as a function of time, at (A) site 1 and (B) site 2. Also shown is the amount of detritus present on day 0 prior to the manipulation and in control plots after 24 weeks (open triangles). The amount of detritus added to DA plots on day 0 is indicated by the open square.

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Table 2 Results of SIMPER analysis, using Bray–Curtis similarity, showing the main taxa that contributed to the cumulative dissimilarity between sites on day 0. The average dissimilarity between sites was 58.8%. The % dissimilarity to standard deviation ratio (Diss/SD) is also given. Taxon

Mean (±SE) abundance (# ind. core−1)

Prionospio aucklandica Aonides trifida Heteromastus filiformis Oligochaeta Austrovenus stutchburyi Arthritica bifurca

23.2 ± 23.3 ± 1.3 ± 4.2 ± 11.5 ± 7.0 ±

Site 1 4.6 1.0 0.5 1.0 1.8 2.8

Diss/SD

Contribution (%)

2.9 5.5 3.2 0.7 1.7 1.0

46.7 14.5 12.1 3.6 3.2 3.1

Site 2 101.0 0.5 20.7 9.7 8.3 5.8

± ± ± ± ± ±

14.1 0.4 3.0 5.0 1.5 1.0

Table 4 p-values of PERMANOVA post-hoc pair-wise comparisons, testing for differences in benthic macrofaunal communities between treatments (C = control, PC = procedural control, DA = detrital addition). Site 1 results are reported above the diagonal and site 2 below (p b 0.05 are indicated in bold).

C C

mangrove detritus has been limited, and comparisons between sites are lacking. Our study explored whether mangrove detritus is an important factor in controlling benthic community structure and variability on temperate intertidal flats. We found that responses of benthic macrofauna to mangrove detrital enrichment were subtle and involved changes in the relative abundances of some species rather than shifts in species composition. Previous detrital addition experiments (on intertidal flats) have concluded that the procedure of the one-off mixing/burial of detrital material does not create variation in macrofaunal assemblages away from an ambient state (Kelaher and Levinton, 2003; Olabarria et al., 2010). However, our results suggest that physical disturbance effects of small-scale surficial sediment mixing may occur in finer sediments with high mud content (site 2), but not in sandy areas (site 1). Although this result does not agree with previous detrital addition experiments (Kelaher and Levinton, 2003; Olabarria et al., 2010), it fits with the literature of physical disturbances on soft-sediment communities. Physical disturbances can create greater responses in muddy sediments compared to sandy sediments (e.g. Ferns et al., 2000; Schratzberger and Warwick, 1999). Our results suggest that the response could differ with sediment properties and this needs to be tested in future studies replicated across multiple site and time scales. Rather than altering benthic species composition, detrital addition (at both sites) resulted in a small decrease in overall macrofaunal abundance, which was largely driven by the reduction in the numerically dominant polychaete species P. aucklandica and to a lesser extent other dominant polychaete species (A. trifida at site 1 and H. filiformis at site 2). Some previous studies examining the effects of macroalgae detritus on intertidal communities have shown similar results, and detected only subtle decreases in a few species following detrital

Table 3 Summary of repeated measures PERMANOVA (Bray–Curtis similarity) comparing benthic macrofaunal community structure between treatment plots (C = control, PC = procedural control, DA = detrital addition) through time (p b 0.05 are indicated in bold), at (A) site 1 and (B) site 2. Source

df

SS

MS

Pseudo-F

p

(A) Site 1 Treatment Time Plot (treatment) Time × treatment Residual

2 4 15 8 59

11,426 7345.5 12,011 6180.1 38,984

5712.8 1836.4 800.7 772.5 660.7

7.1 2.8 1.2 1.2

0.0002 0.0001 0.09 0.19

(B) Site 2 Treatment Time Plot (treatment) Time × treatment Residual

2 4 15 8 60

4475.6 9182.1 7524.6 3591.6 22,983

2237.8 2295.5 501.6 448.9 383.1

4.5 6.0 1.3 1.2

0.0008 0.0001 0.05 0.21

PC 0.2

PC

0.03

DA

0.006

DA 0.002 0.001

0.04

additions. These studies suggested that the intertidal communities examined are accustomed to detrital depositions, and that these systems may be effective at recycling nutrients to buffer the effects of organic enrichment (Olabarria et al., 2010; Rossi, 2006). However, in our system the subtle changes in relative species abundances could be due to the low bioavailability of temperate mangrove detritus (i.e. slow decay and low nutritional quality). Temperate mangrove litter has a high C:N ratio (e.g. 47; Gladstone-Gallagher et al., 2014) compared with other food sources (e.g. microphytobenthos (MPB), C:N of 5–15; Cook et al., 2004; Cook et al., 2009), and is associated with low palatability and nutritional value to benthic consumers (Enríquez et al., 1993; Nordhaus et al., 2011). In addition, temperate litter decay can be an order of magnitude slower than in the tropics (t50 = 63–88 days vs. b 1 week) (Bosire et al., 2005; Gladstone-Gallagher et al., 2014). Therefore, the reduction in C:N content of the litter is also slower, reducing the bioavailability to consumers. Moreover, shallow temperate estuarine sediments often contain a highly productive and palatable MPB (MacIntyre et al., 1996; Miller et al., 1996), meaning that mangrove detritus may not offer the same important food subsidy that has been found in resource-limited tropical systems, such as coral reefs (Granek et al., 2009; Lapointe et al., 1987). Strong community responses to detrital enrichment in temperate estuaries have been observed with additions of macroalgae and/or seagrass (e.g. Bishop et al., 2010; Kelaher and Levinton, 2003; O'Brien et al., 2010), which represent more easily degraded, bioavailable food sources (lower C:N ratios and faster decay rates; Enríquez et al., 1993; de Boer, 2000). The relatively slow decomposition and low bioavailability of temperate mangrove leaf litter may offer one explanation for the weak macrofaunal responses to mangrove detrital deposition. The slow decay rate could also result in minimal changes to sediment biogeochemistry, a factor which drives shifts in species composition (e.g. Kelaher and Levinton, 2003). Temperate mangrove detritus has been found to decrease macrofaunal abundance and diversity at high levels of addition, attributable to the leaching of toxic tannins from leaf litter. However, in lower additions mangrove detritus (and other sources) can have small positive effects on macrofaunal abundance and species richness (Bishop and Kelaher, 2008). The amount of detritus (dry weight) used in our manipulation is similar to the upper end of the range previously added (e.g. Bishop and Kelaher, 2008; Kelaher and Levinton, 2003; Olabarria et al., 2010). In some of these earlier studies, a similar amount of macrophyte detritus created sediment anoxia (during detrital breakdown) instigating negative responses in some macrofauna (Bishop and Kelaher, 2008; Kelaher and Levinton, 2003). However, the large amount of detritus added during this experiment (amount fallen from the trees during summer) did not create anoxic conditions or produce films of sulphide reducing bacteria that have been previously observed with other detrital types (e.g. Bishop and Kelaher, 2008; Kelaher and Levinton, 2003). This comparison further emphasizes the importance

R.V. Gladstone-Gallagher et al. / Journal of Experimental Marine Biology and Ecology 460 (2014) 80–88 Table 5 p-values of PERMANOVA post-hoc pair-wise comparisons, testing for differences in benthic macrofaunal communities between sampling dates (in weeks post addition). Site 1 results are reported above the diagonal and site 2 below (p b 0.05 are indicated in bold).

2 2

4 0.01

8

12

24

0.2

0.0002

0.03

0.6

0.009

0.3

0.004

0.1

4

0.3

8

0.001

0.003

12

0.0001

0.0002

0.0003

24

0.001

0.03

0.1

0.0006 0.0006

of decay rate and suggests that in our system mangrove detrital deposition plays a minor role in shaping benthic macrofaunal communities. The distribution and composition of soft-sediment communities exhibit both temporal and spatial heterogeneity as a result of abiotic and biotic variables (e.g. Morrisey et al., 1992; Thrush, 1991; Thrush et al., 1994). Consequently, responses to the burial of algal wrack have been found to differ among sites with different sediment and benthic community properties (Rossi and Underwood, 2002). Although the sites of the current study had different physical sediment properties, the subtle responses in the relative abundances of dominant polychaetes were the same at both sites. In contrast, the burial of Ulva sp. wrack resulted in species responses that were dependent on initial abundances of species prior to burial, where different species responded on sand flats compared to mud flats (Rossi and Underwood, 2002). Here, both sites responded similarly to the enrichment of mangrove detritus, which

85

could indicate that only a few particular species (e.g. P. aucklandica) are sensitive to organic enrichment and that these responses are not dependent on initial species abundances or sedimentary environment. However, further testing with replication across sedimentary gradients is required. It should be recognised that the statistical power of our experimental design may act as a potential limitation and explanation of the weak response detected. However, the level of replication (n = 6 per treatment) used in our study is comparable to other studies of this nature (n = 3–7) who were able to detect significant detrital effects in macrofaunal communities (e.g. Bishop and Kelaher, 2013a; Rossi and Underwood, 2002). In addition, a previous study conducted in the Whangamata Harbour found that six core samples (13 cm dia. × 15 cm depth) were adequate to detect macrofaunal community variability, and precision of the data was not greatly reduced with six cores compared to twelve (van Houte-Howes et al., 2004). Our design has comparable statistical power to similar studies, which further suggests the absence of a strong benthic community response to mangrove detrital addition. Detrital inputs can also influence lower trophic levels; for example, the response of nematodes to mangrove detritus in tropical sediment communities (Tietjen and Alongi, 1990). Responses of lower trophic levels to detrital deposition could have indirect and cascading effects on higher trophic levels. However, our study did not allow us to detect such changes in lower trophic levels and therefore indirect food web effects, because our sieve mesh size excluded the meiofaunal community of organisms. Furthermore, increases in sediment chlorophyll a content (indicator of MPB biomass) have been observed following detrital enrichment (Bishop and Kelaher, 2007, 2008; Kelaher et al., 2003; Levinton, 1985). The decomposition of plant material can release nutrients, potentially accelerating the growth of micro-organisms such as

Fig. 2. Non-metric MDS ordinations (Bray–Curtis similarity) showing treatment (C = control, PC = procedural control, DA = detrital addition) and time (weeks post addition) effects on benthic community composition at site 1 (A and B, respectively) and at site 2 (C and D, respectively). Each point represents the community in one macrofaunal core sample (13 cm dia.; 15 cm depth).

86

R.V. Gladstone-Gallagher et al. / Journal of Experimental Marine Biology and Ecology 460 (2014) 80–88 Table 6 Results of SIMPER analysis (Bray–Curtis similarity) comparing community composition between treatments (C = control, PC = procedural control, DA = detrital addition) at (A) site 1 and (B) site 2. The table includes the % dissimilarity between treatments, as well as a list of the taxa that contributed N50% to the cumulative dissimilarity. The % dissimilarity to standard deviation ratio (Diss/SD) is also given. Taxon (A) Site 1 C vs. PC Prionospio aucklandica Aonides trifida Austrovenus stutchburyi C vs. DA Prionospio aucklandica Aonides trifida Austrovenus stutchburyi DA vs. PC Aonides trifida Prionospio aucklandica Austrovenus stutchburyi (B) Site 2 C vs. PC Prionospio aucklandica Heteromastus filiformis Oligochaeta C vs. DA Prionospio aucklandica Heteromastus filiformis Oligochaeta DA vs. PC Prionospio aucklandica Heteromastus filiformis Oligochaeta

Fig. 3. Mean (+1 SE) species richness (A) and total abundance (B) at the experimental sites, as a function of treatment (control (C) = open bars, procedural control (PC) = grey bars, detrital addition (DA) = black bars). Means are temporally averaged for each plot to show main effects of treatment (n = 6). At each site, bars not sharing the same letter are significantly different from each other (repeated measures ANOVA, followed by Newman–Keuls; p b 0.05; Table A.1).

Dissimilarity (%)

Contribution (%)

Diss/SD

20.0 17.3 12.8

1.3 1.3 1.3

25.7 21.9 10.6

1.7 1.4 1.3

24.0 19.6 10.3

1.4 1.4 1.2

34.9 10.1 9.0

1.1 1.3 1.0

39.0 9.9 9.6

1.3 1.3 1.0

34.0 13.0 8.5

1.3 1.3 0.8

31.0

44.8

42.7

26.5

31.0

29.2

minor effects on community composition. It is proposed that in these temperate shallow water systems, the highly productive and nutritive microphytobenthos may be a dominant factor controlling spatiotemporal variability of benthic macrofaunal communities (e.g. van der Wal et al., 2008). However, the effects of mangrove detrital deposition may not be important to macrofaunal communities directly, but could influence them indirectly via lower trophic levels. Future studies should endeavour to determine if mangrove (and other macrophytes) detritus affects macrofaunal communities indirectly by fuelling benthic microbial primary productivity.

MPB (Rublee, 1982). We observed no impact of mangrove detrital input on sediment chlorophyll a concentrations, however increases in MPB can be concealed when an effective grazing community is present (Bishop et al., 2007; Levinton, 1985). Despite associated increases in MPB biomass following detrital deposition (Bishop and Kelaher, 2007, 2008; Bishop et al., 2007; Levinton, 1985), no one to date has measured whether this increase in biomass relates to increased production or a relaxation of grazing pressure due to shifts in community structure. Future research should endeavour to determine if mangrove detrital depositions enhance primary productivity of MPB by releasing nutrients during the decomposition process.

5. Conclusions Our results indicate that mangrove detrital deposition plays a relatively minor role in shaping benthic community variability of temperate intertidal flats at the scale tested. Site-specific responses included disturbance effects at site 2, but the same species responded to the detrital addition at both sites. The addition of mangrove detritus did not create shifts in benthic community composition or diversity as hypothesised, but rather caused subtle changes in the relative abundances of a few taxa. The slow decomposition and low nutritional value of temperate mangrove detritus, compared with other detrital sources, could provide an explanation for why a relatively large amount of detritus resulted in

Fig. 4. Mean (+1 SE) abundance of Prionospio aucklandica at the experimental sites, as a function of treatment (control (C) = open bars, procedural control (PC) = grey bars, detrital addition (DA) = black bars). Means are temporally averaged for each plot to show main effects of treatment (n = 6). At each site, bars not sharing the same letter are significantly different from each other (repeated measures ANOVA, followed by Newman– Keuls; p b 0.05; Table A.1).

R.V. Gladstone-Gallagher et al. / Journal of Experimental Marine Biology and Ecology 460 (2014) 80–88

87

Acknowledgements

References

We thank Sarah Hailes, Barry Greenfield, Betty Gubri, Annette Rogers, Kerry Allen, Lisa McCartain, Arie Spyksma, Hannah Jones and Gillian Gladstone for their invaluable help with sample processing; Dudley Bell, Nikki Webb, Simon Brown, Pauline Robert, Daniel Pratt, Jamie Armstrong, Stacey Buchanan, Nicholas Wu, Rachel Harris, Clarisse Niemand, Dorothea Kohlmeier, Kelly Carter, Phil Ross and Marenka Weis for assistance in the field; and Judi Hewitt for statistical advice. We also thank two anonymous reviewers for their constructive comments that greatly improved and simplified the manuscript. This project was funded by a University of Waikato Masters Research Scholarship, the New Zealand Ministry for Science and Innovation Project # CO1X1002, and the Waikato Regional Council.

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Appendix A

Table A.1 Summary of repeated measures ANOVA results comparing treatment (C = control, PC = procedural control, DA = detrital addition) means through time for species richness (A and B), total abundance (C and D) and Prionospio aucklandica (E and F) at sites 1 and 2, respectively. Significant effects are indicated in bold (p b 0.05). Post-hoc Newman–Keuls test results for significant treatment main effects are indicated in Figs. 3 and 4 (p b 0.05). Source of variation

df

Mean-square

(A) Species richness site 1 Intercept 17,112 Treatment 43 Error 90 Time 107 Time × treatment 66 Error 319

SS

F-ratio

p

1 2 15 4 8 60

17,112.0 21.5 6.0 26.7 8.3 5.3

2853.06 3.59

b0.0001 0.05

5.01 1.55

0.0015 0.2

(B) Species richness site 2 Intercept 16,107 Treatment 34 Error 55 Time 68 Time × treatment 34 Error 257

1 2 15 4 8 60

16,106.8 16.9 3.6 17.0 4.2 4.3

4424.96 4.64

b0.0001 0.03

(C) Total abundance site 1 Intercept 2,303,360 Treatment 30,368 Error 18,585 Time 88,505 Time × treatment 9190 Error 62,853

1 2 15 4 8 60

2,303,360.0 15,183.8 1239.0 22,126.4 1148.7 1047.5

1859.07 12.26

b0.0001 0.0007

21.12 1.10

b0.0001 0.4

(D) Total abundance site 2 Intercept 787,251 Treatment 21,791 Error 14,118 Time 10,484 Time × treatment 6061 Error 31,664

1 2 15 4 8 60

787,251.4 10,895.5 941.2 2620.9 757.7 527.7

836.43 11.58

b0.0001 0.0009

(E) Prionospio aucklandica abundance site 1 Intercept 39,021 1 Treatment 5209 2 Error 3262 15 Time 1773 4 Time × treatment 688 8 Error 3767 60

39,020.8 2604.6 217.4 443.2 86.0 62.8

179.46 11.98

b0.0001 0.0008

7.06 1.37

0.0001 0.2

(F) Prionospio aucklandica abundance site 2 Intercept 529,000 1 Treatment 13,584 2 Error 9794 15 Time 9362 4 Time × treatment 4347 8 Error 25,300 60

529,000.0 6792.2 653.0 2340.6 543.3 421.7

810.16 10.40

b0.0001 0.002

5.55 1.29

0.0007 0.3

3.95 0.98

4.97 1.44

0.007 0.5

0.002 0.2

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