DOI: 10.5061/dryad.kd51c5b9f
title: Grazer host density mediates the ability of parasites to protect foundational plants from overgrazing

description: Like many top consumers, parasites can regulate feeding of their prey via
 trait-mediated means. If parasites modify the feeding behavior of
 ecologically important grazers, they may have cascading effects on the
 structure and functioning of whole plant communities. The extent to which
 parasites can influence plant communities in this way is largely dependent
 on the strength of their behavioral alteration, their prevalence in host
 grazers, and the density of those hosts. Recent experiments and
 comparative surveys in southeastern USA salt marshes revealed that common
 larval trematode parasites suppress the per capita grazing impacts of the
 marsh periwinkle (Littoraria irrorata), generating a trophic cascade that
 protects foundational marsh plants from drought-associated overgrazing.
 Here, we conducted a field manipulation wherein we modified grazer host
 density while holding infection prevalence constant at an ecologically
 relevant level (20%) to determine whether the indirect, facilitative
 effects of parasites on marsh plants varied with the density of grazers.
 We found that parasites had significant positive impacts on marsh net
 primary productivity at moderate densities of snails (≥50 snails/ 0.5 m2),
 but that the positive effects of parasites were negligible at lower
 densities. Our results confirm the findings of previous studies that
 parasites can protect marsh plants from overgrazing at sufficiently high
 prevalence but show that their ability to do so depends on host density.
description: To determine how
 trematode infection influences marsh productivity at different levels of
 grazer host density, we conducted a field manipulation where we modified
 grazer host density while keeping infection prevalence constant. In June
 2016 we established 0.5 m<sup>2</sup> caged plots in a
 structurally homogenous swath of smooth cordgrass
 (<em>Spartina</em>
 <em>alterniflora</em>)<em> </em>marsh within the
 Hoop Pole Creek Clean Water Reserve in Atlantic Beach, North Carolina, USA
 (34°42’25.12” N, 76°45’1.14” W). The site was characterized by a
 relatively uniform elevation, very low snail densities (&lt;1 adult
 snail per m<sup>2</sup>), and minimal visible signs of snail
 grazing. Snails used in the experiment
 were collected from a marsh die-off area where snails were abundant and
 infection prevalence was known to be high (Morton & Silliman
 2020). Collected adult snails (shell length &gt; 15mm) were
 transported back to the lab where their infection status was determined
 using a previously described cercariae shedding method that produces no
 false-positives (Morton 2018, Morton & Silliman 2020). We marked
 the shell of each infected snail with a red dot using a non-toxic,
 water-resistant paint pen while uninfected snails were marked with a blue
 dot (Henry & Jarne 2007). Snails were kept in separate aquaria and
 provisioned with damp cordgrass<em> </em>wrack for ~3 weeks
 until they were deployed in the field. Roofless cages (0.7 × 0.7 × 1
 m) were constructed from untreated wooden posts and galvanized hardware
 mesh. A strip of copper tape was applied to the inner base of each cage,
 just above the sediment, to discourage snail escapes. Caged plots were
 spaced at least 1-m apart to assure independence of replicates—a design
 confirmed from past studies (Silliman & Zieman 2001, Silliman
 & Bertness 2002, Morton & Silliman 2020). Cages were
 buried 10 cm into the substrate to prevent snail and mud crab migration in
 and out of cages, and to inhibit belowground connections between plants
 inside and outside of the cages. Each plot was assigned to one of 8 snail
 density treatments (20, 40, 50, 60, 70, 80, 90, and 100 snails/per 0.5
 m<sup>2</sup>) and one of two parasite addition treatments (0
 and 20% infection prevalence). This resulted in 16 total treatments (n = 4
 replicates per treatment). The snail densities used spanned the full range
 of adult snail densities observed within marsh die-off areas at this site.
 The 20% infection prevalence used in the experiment reflected the average
 naturally occurring summertime infection prevalence value for snails
 within local die-off areas (Morton & Silliman 2020). Uncaged plots
 marked at the corners with colored PVC flags (n = 4) and partial cages
 with one open side (n = 4) served as cage controls.
 Before the beginning of
 the experiment, we removed any mud crabs and snails from cages. We took
 measurements of several marsh characteristics in all plots at the
 beginning of the experiment. We counted all <em>S.
 alterniflora</em> stems and measured the heights of ten randomly
 selected stems in each plot. Random selection of stems was accomplished by
 tossing a plastic dowel into plots and measuring the first 10 stems
 touching the dowel. We constructed a height-to-biomass regression by
 collecting 30 cordgrass stems of varying sizes from the marsh directly
 adjacent to our experimental site. The stems were washed, their length
 measured, and were dried at 70˚C until they reached a constant weight. We
 used the resulting height-to-biomass regression to estimate standing
 cordgrass biomass in each plot. We also counted all juvenile (&lt; 0.5
 cm in diameter) and adult fiddler crab burrows at both the beginning and
 end of the experiment, because these organisms are known to influence
 cordgrass growth by oxygenating the sediment through burrowing (Bertness
 1985, Daleo et al. 2007, Angelini & Silliman 2012, Gittman
 & Keller 2013, Raposa et al. 2018). Infected and uninfected snails
 within each plot were counted twice weekly and replaced as necessary to
 maintain the assigned snail density and infection prevalence treatments
 for the 3-month duration of the experiment. During monitoring, any
 predatory mud crabs found within plots were removed and their burrows
 plugged with marsh sediment to discourage successful
 re-occupation. We took final metrics of marsh
 vegetation characteristics in September 2016, 12 weeks after the beginning
 of the experiment. All snails were removed and dissected to confirm
 infection status. Because snail grazing had dramatically reduced stem
 densities in many plots, we were able to measure all stems within each
 plot to generate final plot biomass estimates by calculating and summing
 individual biomass estimates. Belowground biomass cores (15-cm diameter x
 25-cm height) were also taken from the center of plots. Cores were washed
 and root material separated from rhizomes. Roots and rhizomes were dried
 to a constant weight at 70˚C and weighed. The possible influence of
 caging artifacts on marsh invertebrates and cordgrass growth was assessed
 using paired t-tests. Specifically, we compared the number of total
 fiddler crab burrows, final stem density, initial aboveground biomass,
 final aboveground biomass, change in aboveground biomass, and number of
 flowering stems between cage controls and open controls.
 Because fiddler crabs can
 positively influence cordgrass growth (Bertness 1985, Gittman &
 Keller 2013), we examined the relationship between burrow density and
 cordgrass stem density using a linear model. We validated model fit by
 examining model residuals, their distribution relative to fitted values,
 and normal Q-Q plots (<em>car</em> package, Chambers and
 Hastie 1992). In accordance with our
 hypotheses, models including the interaction between parasite infection
 (uninfected versus infected) and snail density (20, 40, 50, 60, 70, 80,
 90, or 100 snails per 0.5 m<sup>2</sup>) were fit for each
 response metric (belowground biomass, aboveground biomass, shoot density,
 and number of flowering stems). Additionally, all models initially
 included fiddler crab burrow density as a covariate. However, this
 covariate was ultimately not significant for any model and was therefore
 dropped and each model re-run to include only the interaction term. We
 analyzed the interactive effects of infection status and snail density on
 belowground biomass using a two-way analysis of variance (ANOVA), with
 data log-transformed prior to analysis to meet model assumptions. Change
 in aboveground biomass was analyzed with a linear model, with initial
 aboveground biomass values for each plot subtracted from corresponding
 final values. Model assumptions were verified through assessments of
 homogeneity of variance (Levene’s test, <em>P</em> &gt;
 0.05) and examination of fitted residuals and normal Q-Q plots. The number
 of cordgrass inflorescences was modeled with a negative binomial
 generalized linear model (<em>glmmTMB</em>, Brooks et al.
 2017). Model appropriateness and fit were confirmed through examination of
 simulated scaled residuals (<em>DHARMa</em>, Hartig 2022).
 For each response metric, the
 significance of the interaction model was examined in an Analysis of
 Deviance Table using Wald chi-square tests (<em>car
 </em>package, Chambers & Hastie 1992). For all models,
 Tukey’s post-hoc comparisons were used to assess pairwise differences for
 any significant treatment or interactive effects in the models
 (<em>emmeans </em>package, Lenth et al. 2021). All analyses
 were performed in the R statistical computing environment (v. 4.1.3; R
 Core Team 2018).