# DEATH AND DECOMPOSITION IN AQUATIC ECOSYSTEMS

EDITED BY : M. Eric Benbow and Gary A. Lamberti PUBLISHED IN : Frontiers in Ecology and Evolution

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# DEATH AND DECOMPOSITION IN AQUATIC ECOSYSTEMS

Topic Editors: M. Eric Benbow, Michigan State University, United States Gary A. Lamberti, University of Notre Dame, United States

Citation: Benbow, M. E., Lamberti, G. A., eds. (2021). Death and Decomposition in Aquatic Ecosystems. Lausanne: Frontiers Media SA. doi: 10.3389/978-2-88966-339-2

# Table of Contents

*04 Death and Decomposition in Aquatic Ecosystems* M. Eric Benbow, Joseph P. Receveur and Gary A. Lamberti *16 Quantitative Food Webs Indicate Modest Increases in the Transfer of Allochthonous and Autochthonous C to Macroinvertebrates Following a Large Wood Addition to a Temperate Headwater Stream* Sally A. Entrekin, Emma J. Rosi, Jennifer L. Tank, Timothy J. Hoellein and Gary A. Lamberti *32 Drought-Induced, Punctuated Loss of Freshwater Mussels Alters Ecosystem Function Across Temporal Scales* Traci P. DuBose, Carla L. Atkinson, Caryn C. Vaughn and Stephen W. Golladay *45 Mass Mortality Events of Invasive Freshwater Bivalves: Current Understanding and Potential Directions for Future Research* William G. McDowell and Ronaldo Sousa *57 Fish, Including Their Carcasses, Are Net Nutrient Sources to the Water Column of a Eutrophic Lake* Regina L. G. Nobre, Luciana S. Carneiro, Sarah E. Panek, Maria J. González and Michael J. Vanni *66 Microbial Community Response to a Novel Salmon Resource Subsidy* Courtney E. Larson, Jennifer L. Pechal, Brandon S. Gerig, Dominic T. Chaloner, Gary A. Lamberti and M. Eric Benbow *80 Landscape Structure and Species Interactions Drive the Distribution of Salmon Carcasses in Coastal Watersheds* Joel M. S. Harding, Jennifer N. Harding, Rachel D. Field, Jane E. Pendray, Noel R. Swain, Marlene A. Wagner and John D. Reynolds *95 Understanding the Relative Roles of Salmon Spawner Enrichment and Disturbance: A High-Frequency, Multi-Habitat Field and Modeling Approach*

Janine Rüegg, Dominic T. Chaloner, Ford Ballantyne, Peter S. Levi, Chao Song, Jennifer L. Tank, Scott D. Tiegs and Gary A. Lamberti

Amanda L. Subalusky, Christopher L. Dutton, Emma J. Rosi, Linda M. Puth and David M. Post

*134 Editorial: Animal Mass Mortalities in Aquatic Ecosystems: How Common and Influential?*

Gary A. Lamberti, Natalie M. Levesque, Michael A. Brueseke, Dominic T. Chaloner and M. Eric Benbow

# Death and Decomposition in Aquatic Ecosystems

#### M. Eric Benbow1,2,3 \* † , Joseph P. Receveur1 and Gary A. Lamberti4

 Department of Entomology, College of Agriculture and Natural Resources, Michigan State University, East Lansing, MI, United States, 2 Department of Osteopathic Medical Specialties, Michigan State University, East Lansing, MI, United States, Ecology, Evolutionary Biology and Behavior Program, Michigan State University, East Lansing, MI, United States, Department of Biological Sciences, College of Science, University of Notre Dame, Notre Dame, IN, United States

Resource subsidies affect nutrient cycling, species interactions, and food webs in ways that influence ecosystem structure and function, but their effects depend on the history, magnitude, and recurrence frequency of the subsidies. In aquatic ecosystems, plant detritus has been considered the predominant form of such subsidies; however, while considered less abundant in many ecosystems, carrion represents subsidies with relatively rapid turnover and highly concentrated nutrient and energy release that can have strong and lasting effects on ecosystems. Carrion subsidies can be both autochthonous or allochthonous, and come in the form of natural senescence or disease-related non-consumptive mortality, phenology-based programed death (e.g., salmon spawning and death), or stochastic and episodic events (e.g., mass fish die-offs). All aquatic ecosystems have some level of non-consumptive mortality that provides a background level of carcasses to aquatic ecosystems, while others have a natural history of carrion resource subsidies (e.g., natural salmon-bearing streams), and some have only recently been exposed to phenology-based carrion subsidies (e.g., anthropogenic salmon introductions around the world). Many aquatic ecosystems experience episodic subsidies in the form of unexpected mass mortalities (e.g., eutrophication-, disease-, or climate-related mass die-offs) or have seasonally dependent pulses, like that of marine or lake snow in the form of zooplankton carcasses. The responses of ecosystems to these different histories and frequencies of carrion subsidies have often been independently investigated, with little effort to compare and bridge research boundaries in the broader context of resource subsidies. In this review, we provide a synthesis of how pulsed carrion nutrient and energy subsidies have widespread and lasting impacts on many aquatic ecosystems. We do this with a synthesis of literature from freshwater and marine ecosystems along three themes of how carrion is produced and decomposes: autochthonous and allochthonous necromass; phenology-based mortality; and stochastic and episodic mass mortality subsidies. Studies of charismatic megafauna carrion (e.g., whales) have described significant impacts in deep ocean systems, but much less is understood for other groups of animals. Quantifying the energy, nutrient, and foodweb effects of carrion is needed for more species and among habitats to more fully understand how ubiquitous forms of necromass contribute to aquatic ecosystem structure and function.

Keywords: decomposition, microbial, ecosystem metabolism, community assembly, necrobiome, carrion, mass mortality, forensics

#### Edited by:

Alexei B. Ryabov, University of Oldenburg, Germany

#### Reviewed by:

Ralf Schulz, University of Koblenz Landau, Germany Zhang Yixin, Soochow University, China

> \*Correspondence: M. Eric Benbow<EMAIL_ADDRESS>
†ORCID: M. Eric Benbow orcid.org/0000-0003-2630-0282

#### Specialty section:

This article was submitted to Population and Evolutionary Dynamics, a section of the journal Frontiers in Ecology and Evolution

Received: 29 August 2019 Accepted: 21 January 2020 Published: 07 February 2020

#### Citation:

Benbow ME, Receveur JP and Lamberti GA (2020) Death and Decomposition in Aquatic Ecosystems. Front. Ecol. Evol. 8:17. doi: 10.3389/fevo.2020.00017

**4**

### INTRODUCTION TO NECROMASS RESOURCE SUBSIDIES IN AQUATIC ECOSYSTEMS

#### Background

Decaying organic matter, or necromass, comes in the form of both plant and animal biomass, recently differentiated as autotrophically (e.g., leaf litter) and heterotrophically (e.g., carrion) derived biomass, respectively, and is recycled back into ecosystems by the necrobiome community (Benbow et al., 2019). While both forms of necromass are considered important to ecosystems, plant detritus processing has historically dominated research as the main energy pathway in aquatic ecosystems (Vannote et al., 1980; Webster and Benfield, 1986; Getz, 2011; Boyero et al., 2016), especially for lotic systems lacking anadromous fish populations (Moore et al., 2004; Gessner et al., 2010; Benbow et al., 2019). Until recently (McDowell et al., 2017; Subalusky et al., 2017; DuBose et al., 2019; Wenger et al., 2019), aquatic studies of animal necromass (carrion) have largely focused on fish carcass decomposition (Richey et al., 1975; Garman, 1992; Schindler, 1992), or lake/marine snow represented by zooplankton carcasses in the water column of lentic ecosystems (Alldredge and Silver, 1988; Grossart and Simon, 1998; Giering et al., 2014). Some research has addressed the long-term decomposition of whale (and other vertebrate) falls in the oceanic abyss (Allison et al., 1991; Bennett et al., 1994; Smith and Baco, 2003; Kemp et al., 2006; Higgs et al., 2014). Much of past research on carrion in aquatic ecosystems has focused on species that are large in size, have anthropogenic importance (e.g., salmon runs), or elicit public intrigue (e.g., whale falls); however, a large portion of aquatic ecosystem necromass comes in the form of smaller organisms (e.g., phytoplankton, zooplankton, and invertebrates) with fast generation times and high turnover rates that substantially contribute to ecosystem production (Waters, 1977; Benke, 1998; Huryn and Wallace, 2000; Landry and Calbet, 2004; Patrick et al., 2019). While we acknowledge the importance of phototrophically derived necromass in aquatic ecosystems and the associated trophic relationships (Little and Altermatt, 2018), we direct interested readers to the many extensive reviews and empirical research of such resources ranging from phytoplankton, seaweed, and macrophytes to large wood debris (Lindeman, 1942; Mann, 1969; Wallace et al., 1999; Moore et al., 2004; Entrekin et al., 2009; Tank et al., 2010).

For this review, we focus on heterotrophically derived necromass (from bacteria to whales) with emphasis on animal carrion, and also acknowledge that dung, frass, and other forms of animal tissue (e.g., gametes) contribute to the larger resource pool (Subalusky et al., 2015; Dutton et al., 2018; Subalusky and Post, 2018; Benbow et al., 2019). For an informative argument for the importance of egested heterotrophically derived forms of particulate organic matter to energy budgets of pelagic zones of lake ecosystems we direct readers to Wetzel (1995).

In general, the importance of carrion has been qualitatively considered relatively greater in large lakes and oceans (Alldredge and Silver, 1988; Smith and Baco, 2003; Tang et al., 2014) compared to shallow wetlands and marshes (Brinson et al., 1981; Duggins et al., 1989). However, Subalusky and Post (2018) discuss how recipient ecosystem qualities (e.g., productivity or trophic structure) are important for understanding crosssystem effects of carrion. In watersheds, carrion has been largely studied as pulsed resource subsidies like fish die-offs (Parmenter and Lamarra, 1991) or as part of programed phenologically driven semelparous death like salmon life cycles (Moore et al., 2004; Benbow et al., 2019). Regardless of origin, most organic matter ultimately decomposes and is recycled in ecosystems (Wetzel, 1995; Moore et al., 2004; Benbow et al., 2019), providing an intimate connection between ecosystem structure and function through interacting necrobiome species responsible for the breakdown (physical destruction into smaller and smaller units) and decomposition (biochemical alteration and conversion) of necromass (Lindeman, 1942; Putman, 1983). This process of decomposition, or turnover rate, is also known to be much faster for carrion compared to plant litter biomass, and is more important to ecosystems than once considered (Barton et al., 2019).

### Precedent From Terrestrial Ecosystems

Most studies of carrion ecology have been conducted in terrestrial ecosystems. Here we provide examples from terrestrial habitats to show the potential of carrion to ecosystems, which can potentially be transferred to aquatic conditions. In terrestrial systems, carrion has often been thought to contribute marginally to ecosystem energetics (Swift et al., 1979; Barton et al., 2019). However, Barton et al. (2019) argue that because of the high turnover rate of carrion, carcasses have historically gone understudied since they are quickly recycled back into the ecosystem and are hidden from observation; thus, they have been presumed to have a negligible contribution to energy and nutrient flow in ecosystems. This view is plausible because it has been historically difficult to quantify natural rates of carrion decomposition in ecosystems, and so the relative production of energy from carrion compared to the production from an equivalent amount of plant necromass may be disproportionate and underestimated (Barton et al., 2019). As one example (see review by Scott, 1998), burying beetles have been shown to remove and conceal 91% of exposed small (21–210 g) mammal carcasses by burial within an average of 1.4 days (Trumbo, 1992). In the same study, 22.7% of the exposed carcasses were eaten or removed by vertebrate scavengers. This example demonstrates how quickly small carrion can be removed from scientists' observation.

Decomposing carcasses contribute hot spots of nutrient release (Carter et al., 2007; Benbow et al., 2019) that in terrestrial systems can have direct and indirect long-term (e.g., months to years) effects on soil conditions (Bornemissza, 1957; Strickland and Wickings, 2015), plant communities (Towne, 2000; Wardle et al., 2004), and both invertebrate and vertebrate scavengers (Bump et al., 2009; Beasley et al., 2012; Barton et al., 2013; Benbow et al., 2015). The spatial extent of such effects in ecosystems has not been well documented, but could be potentially important over decades (Bump et al., 2009). As an example of this potential, Hawlena et al. (2012) found

that fear of spider predation by living grasshoppers resulted in carcasses with significantly higher carbon to nitrogen ratios than non-stressed specimens, and the resulting change in carcass quality, even when the biomass was about 140 times lower than the plant litter biomass, affected below ground community function and subsequent leaf litter decomposition on carcass sites in an old prairie ecosystem. Thus, even small amounts of carrion biomass have significant effects on measurable ecosystem processes, such as leaf litter decomposition and soil function, suggesting that if scaled by population density, mortality rates, and turnover in a landscape carcasses may significantly impact ecosystem functions in ways yet to be examined. If scaled in this manner, carrion will likely be shown to have collective ecosystem level effects like those recognized for ungulate dung and urine deposition in prairies (Norman and Green, 1958; Seastedt et al., 1991) or hippopotamus dung for some African rivers (Subalusky et al., 2015).

#### The Need to Scale From Carcass to Collective Effects

The hypothesis that individual carcass effects can have large spatial and long temporal scale effects has been supported by work in Isle Royale National Park, Michigan (Bump et al., 2009). In this 50-year study of the effects of over 3,600 wolf-killed moose carcasses on landscape heterogeneity and ecosystem function, Bump et al. (2009) reported that soil nutrients (e.g., 100–600% higher inorganic nitrogen at carcass compared to control sites), microbial biomass, microbial community composition, and surrounding plant leaf nitrogen (e.g., leaf nitrogen was 25–47% greater at carcass sites) was significantly elevated for at least 2–3 years at carcass deposition locations. In case studies of whale falls, Smith and Baco (2003) estimate that the sediment beneath whale carcasses receive a pulse of organic matter that is equal to almost 2000 years of background material over decomposition. Furthermore, in observations on large elasmobranch (i.e., whale shark and mobulid rays) carcass falls, Higgs et al. (2014) estimated that such carcasses represent on average about 4% of the normal particulate organic carbon flux to the seafloor in the bounding area of their occurrence. The authors suggest that the deep-sea scavenger communities benefit most from these significant energetic subsidies.

The same considerations for the underappreciated role of carrion in terrestrial ecosystems or the deep sea abyss is true for other aquatic ecosystems, where studies have documented localized carcass effects in salmon-bearing streams, whale falls, and other fauna of the deep ocean benthos (Smith and Baco, 2003; Kemp et al., 2006; Anderson and Bell, 2014; Higgs et al., 2014), including prehistoric carrion (Reisdorf et al., 2012; Danise et al., 2014); fish and waterfowl carcasses of salt marshes (Parmenter and Lamarra, 1991); and additional carcass effects likely in most aquatic habitats.

Indeed, vertebrate carrion placed in deep (300–3000 m) marine habitats initiates a succession of both invertebrate and vertebrate scavengers that take advantage of this punctuated, heterotrophically derived subsidy (Kemp et al., 2006; Anderson and Bell, 2016). What is less understood is how individual carcasses scale with mortality at population and community levels of biological organization. New studies are needed to better use population demographics and mortality estimates for calculating ecosystem level carrion production from species that do not have phenology-based or episodic mass mortalities. In Figure 1 of Barton et al. (2019) aquatic carrion biomass ranged from 102 kg/km2 for copepods up to 107 kg/km2 for bivalves, with different species of fish ranging from about 103 to 106 kg/km2 . Among the examples they provided, terrestrial vertebrate biomass was at least two orders of magnitude lower than aquatic species at 103 kg/km2 . While this example is not comprehensive, it does suggest that the effect of heterotrophically derived necromass in ecosystems is likely much greater than historically considered. To better incorporate the contributions of carrion in aquatic ecosystem energy and nutrient budgets, additional studies are needed that directly quantify carcass necromass and turnover across biomes. Such broadly available data on non-consumptive heterotrophically derived necromass in ecosystems would advance theory in consumer-resource and food web ecology (Getz, 2011).

In this paper we support the arguments of Subalusky and Post (2018) and Barton et al. (2019) that carrion resources are important subsidies in most ecosystems. We do this with a synthesis of literature from freshwater and marine ecosystems along three themes of how carrion is produced and decomposes through the following: autochthonous and allochthonous necromass decomposition; phenology-based mortality; and stochastic and episodic mass mortality subsidies. We provide examples and case studies that collectively show that carrion, along a size continuum that spans multiple orders of magnitude, likely plays an underappreciated role in ecosystem energetics and nutrient dynamics. Furthermore, we posit that death links living resource pools of multiple and interacting populations of species through pathways of nonconsumptive mortality that result in a dead resource pool represented by carcasses that are quickly decomposed and recycled back into the living resource pool through detritivores and omnivores (**Figure 1**). In aquatic systems, such carrion represents autochthonous or allochthonous energy and nutrient subsidies that affect foodwebs and ecosystem function in complex and understudied ways.

### AUTOCHTHONOUS AND ALLOCHTHONOUS CARRION NECROMASS

Similar to phototrophically derived detritus, carrion necromass in aquatic ecosystems can come in the form of autochthonous and allochthonous sources. Carrion necromass of aquatic ecosystems can be generated from within or outside of a system in the form of autochthonous and allochthonous resources, respectively (**Figure 1**). As part of the autochthonous resource pool, carcasses resulting from natural senescence, physiological intolerance, or disease-related death of aquatic heterotrophs enter the detrital pool based on rates of mortality, disease, partial predation, and generation times related to natural senescence. These resources

represent necromass from all organisms that live in aquatic habitats, from bacteria to blue whales (Minshall et al., 1991; Fenoglio et al., 2005; Subalusky and Post, 2018).

For any specific taxon, the degree of natural senescence, physiological intolerance (e.g., temperature, physical and chemical thresholds), or disease-related non-consumptive death contribute to the within-system detrital pool and associated energy and nutrient dynamics. For instance, non-consumptive mortality is a natural part of the population dynamics of any species, but the overall magnitude is rarely quantified, except for studies related to aquaculture production of commercially important taxa (Rowe et al., 1989; Karunasagar et al., 1994; Chen et al., 1995; Lorenzen, 1996). Furthermore, the mere presence of predators is known to have negative effects on life history traits and fitness of many species (Preisser et al., 2005). Since many organisms are prey for a variety of predators, this collective effect of non-consumptive predation threats can lead to facilitated senescence within many animal populations, although this form of non-consumptive mortality has not been broadly studied. In general, non-consumptive predator effects have been documented to be as strong or stronger than direct predation effects for many species, and the non-consumptive predator presence effect is generally stronger in aquatic than terrestrial ecosystems (Preisser et al., 2005).

Non-consumptive effects can come in the form of increased costs of defensive strategies that include lower mating success, increased vulnerability to other predators, energetic investments related to finding resources or defensive structures (e.g., morphology or biochemical), or through reduced survivorship (Kotler et al., 1993; Preisser et al., 2005; Sheriff et al., 2009). For instance, McCauley et al. (2011) reported significant non-consumptive predator induced mortality and failed metamorphosis to adults in dragonfly larvae exposed to predatory fish and other dragonfly predators. Depending on predator density, survivorship was 1.2–4.3 higher under nopredator mesocosm conditions. The mechanisms responsible for such non-consumptive predator induced mortality are not clear, but have been speculated to be related to increased susceptibility to disease and energetic costs of foraging related to induced stress (Ramirez and Snyder, 2009; Hawlena and Schmitz, 2010). This example demonstrates that just the presence of a predator leads to greater mortality, supporting the hypothesis that there are degrees of non-consumptive mortality that occur in natural environments; however, the mechanisms and sources of

non-consumptive mortality are difficult to identity and quantify. Such non-consumptive mortality may result in increased heterotrophically derived necromass contributions to ecosystem organic matter budgets through decomposition (**Figure 1**).

The non-consumptive component of population mortality has not been studied in any detail for most aquatic organisms, and is often presumed negligible in ecosystem level budgets or calculations of secondary production where predator consumption rates have often been assumed to lead to 100% prey mortality (Hynes, 1970; Waters, 1977; Wetzel, 1995). This absolute consumption rate is not often the case in natural systems, with many individuals of a population avoiding predation and succumbing to other means (e.g., disease or starvation) of mortality (Wetzel, 1995). Thus, the availability of carcasses to aquatic ecosystems is likely larger than previously assumed, especially if the effects of individual carcasses are scaled by their collective density and rates of availability in the environment as discussed by Barton et al. (2019).

The effect of individual carcasses on ecosystem structure and function has been shown to be variable (Minshall et al., 1991; Fenoglio et al., 2005; Barton et al., 2019; Benbow et al., 2019); however, individual carcass effects have not been scaled to account for population level mortality over space and time in a way that would reveal the collective pool of heterotrophically derived necromass for ecosystems based on natural, nonconsumptive rates of mortality (Barton et al., 2019). A significant challenge to addressing this need lies with differentiating natural senescence, starvation, climate, or disease-related mortality from predation or consumption rates for a specific population. Quantifying non-consumptive mortality is inherently difficult, but could potentially be done by evaluating life tables (Deevey, 1947) of organisms with and without predators, much like that for humans (Haldane, 1953). Additional studies are needed to devise ways to account for non-consumptive mortality, much like that of recent examples that quantified seal (Quaggiotto et al., 2018) and wildebeest (Subalusky et al., 2017) mortality on aquatic ecosystems. However, there have been surprisingly few studies on non-consumptive mortality of micro- and macroinvertebrates; those taxa that can often have fast generation times and high mortality.

Potential ways to quantify non-consumptive mortality of macroinvertebrates can come from life history and secondary production studies in the absence of predators in natural ecosystems or from more artificial conditions, such as those conducted for ecotoxicology studies and aquaculture. Some estimates of macroinvertebrate secondary production are available from fish-bearing and fishless lakes (Arnott and Vanni, 1993; Northington et al., 2010), but the body size distributions are not often reported to allow for estimates of non-consumptive mortality and how that may mediate secondary production. It is often assumed in studies of secondary production that population loss can all be attributed to predation, but this is likely not the case. Data from control groups (e.g., no treatment with a contaminant) used in ecotoxicology studies may also provide data on non-consumptive mortality, but often these experiments are done under artificial conditions associated with laboratory or field mesocosms, where densities and abiotic conditions may not represent natural conditions (Rand et al., 1995; Boudou, 2018). These artificial conditions are similar to aquaculture systems and associated research (Huet et al., 1986). The degree to which nonconsumptive mortality in aquaculture conditions differs from natural ecosystems is not well understood. Another potential mechanism for determining non-consumptive mortality in aquatic populations, and employed in fisheries management, is to derive mortality–weight relationships in populations in different natural ecosystems compared to conditions with eliminated or significantly reduced predation pressure (Lorenzen, 1996, 2000). For instance, Lorenzen (1996) reported allometric scaling of fish mortality to non-predatory mortality by modeling mortality–weight relationships of fish populations from natural ecosystems compared to ponds/cages and tanks with no predation pressure. In these conditions, mortality is attributed to diseases, water quality problems, or winter starvation (Huet et al., 1986). The derived weight exponents of mortality were consistently negative for populations in ponds/cages/tanks compared to natural ecosystems, suggesting non-predatory mortality is more weight (i.e., surrogate for age) dependent than is predatory mortality.

Additional information on carrion impacts on aquatic ecosystems can be gathered from mass mortality studies of autochthonous heterotrophs (see the section "Stochastic and Episodic Mass Mortality and Decomposition") and those related to programed phenology-based death of allochthonous taxa. As part of allochthonous resources, carcasses ultimately come from outside of the system and may include anadromous (e.g., salmon, sturgeon) or catadromous (e.g., eels) vertebrates, crustaceans, and molluscs (e.g., amphidromous shrimp and snails) that spend a portion of their life cycle growing and developing in other habitats or ecosystems (e.g., ocean or mangroves), but complete their life cycle in the freshwater environment (McDowall, 1988; Cederholm et al., 1989; Thuesen et al., 2011; Weaver et al., 2018). These resource subsidies can be made available through natural senescence, physiological intolerance, starvation, or disease-related, non-consumptive death (e.g., amphidromous herring) or through programed phenology-based mortality (e.g., post-spawning salmon). The effects of allochthonous sources of necromass have been studied in considerable detail, especially for salmon, both within its native range (Cederholm et al., 1989; Schindler, 1992; Chaloner et al., 2002; Janetski et al., 2009), but also where it has been introduced to naïve watersheds (Richey et al., 1975; Schuldt and Hershey, 1995; O'Toole et al., 2006).

Parasitized terrestrial insects can also represent allochthonous inputs in aquatic systems, entering streams from the riparian canopy due to modified behavior associated with their parasites (Schmidt-Rhaesa, 2001; Thomas et al., 2002). In one well-documented example, the horsehair worm (Nematomorpha) infects the camel cricket (Orthoptera: Rhaphidophoroidae) in Japanese watersheds. Once infected, the riparian crickets either slowly enter or jump into the streams, upon which the parasite leaves the body immediately or soon after the cricket is dead. These terrestrial subsidies have been shown to lower predation of resident aquatic invertebrates (Sato et al., 2008, 2011a,b, 2012). While not

as well studied, other parasite-infected terrestrial insects (e.g., praying mantis) can also enter water with immediate horsehair release (Schmidt-Rhaesa, 2001). These examples suggest that parasitized insect subsidies enter the detrital pool upon death if not consumed, and may also offer a predatory release of other species that could potentially lead to additional non-consumptive mortality of those species in stream ecosystems.

#### PHENOLOGY-BASED MORTALITY AND DECOMPOSITION

A variety of aquatic organisms use phenology-based cues, such as temperature and day length, to control important behaviors including migration, hatching, and spawning. While these behaviors play critical roles in the fitness of the organisms themselves, they also affect other taxa throughout the food web. For example, the movement of anadromous fish can transport large amounts of nutrients and biomass upstream through their spawning activities and phenology-based mortality (Cederholm et al., 1999; Schindler et al., 2003; Wipfli et al., 2003), which are then used by other consumers (Bilby et al., 1996; Chaloner et al., 2002; Baxter et al., 2005; Hocking and Reynolds, 2011). Although mass mortality events of semelparous fish (e.g., salmon) following spawning are a dramatic introduction of necromass into freshwater systems, the release of eggs (which commonly exhibit high rates of mortality) and excrement by both iteroparous and semelparous fish comprise an important, though less studied, component of the available necromass, in some cases exceeding the nutrient inputs from carcasses (Tiegs et al., 2011; Childress and McIntyre, 2015). These inputs represent a major linkage of marine and freshwater systems as many anadromous fish derive most of their mass (>95%) from marine-based sources before migrating into freshwater systems via migration and subsequent semelparous death (Mathisen et al., 1988; Cederholm et al., 1999; Lamberti et al., 2010).

In contrast to stochastic mass mortality events, phenologybased events provide a regular input of nutrients into aquatic systems, with many organisms altering their life histories to coincide with these predictable influxes of resources (Hocking and Reynolds, 2011; Lisi and Schindler, 2011; Deacy et al., 2017). Predators, such as mink (Mustela vison) or brown bears (Ursus arctos), alter their behavior and timing of reproduction to coincide with the availability of salmon runs (Ben-David, 1997). These scavengers use salmon as a major nutritional source, in some locations obtaining >90% of their carbon and nitrogen from these fish (Willson and Halupka, 1995; Hilderbrand et al., 1996, 1999). The landscape can also affect how salmon carcass resource subsidies impact terrestrial and estuarine ecosystems, with differences reported in how wolves and bears transport carcasses to riparian forests and meadows depending on stream size and location within the watershed (Harding et al., 2019). A range of other mammals, birds, and insects use the carcasses of anadromous fish as resources that can affect the decomposition dynamics in streams (Cederholm et al., 1989; Zhang et al., 2003) and alter their behavior to better use and consume these resources (Moore and Schindler, 2010). In addition to direct consumption by eukaryotes, carcasses in aquatic systems can have strong interactions with microbial communities (Wipfli et al., 1998; Pechal and Benbow, 2016; Pechal et al., 2019). Although the impact of decomposing carcasses is highly dependent on biotic and abiotic factors (e.g., stream physical structure, riparian conditions, organisms present, etc.) the input of nutrients from carcasses stimulates microbial activity and primary production (Mitchell and Lamberti, 2005), leading to additional effects on higher trophic levels (Wipfli et al., 1998; Cederholm et al., 1999; Gende et al., 2002).

Salmon carcasses can also act as a resource subsidy to estuaries of salmon-bearing streams (Cederholm et al., 1999; Gende et al., 2004; Cak et al., 2008), with linkages to macroalgae through copepods (Fujiwara and Highsmith, 1997), marine invertebrates including echinoderms and crustaceans (Reimchen, 1994), and vertebrate taxa, such as coyotes and wolves (Gende et al., 2004). The availability of these resources to estuary systems can be mediated by complex interactions among trophic groups and habitat conditions. While the feeding activity of gray wolves (Canis lupus) and bears (Ursus spp.) can transfer carcasses from stream reaches to riparian habitats (Gende et al., 2004), where they become available to other scavengers, how many and which species of salmon they transfer depends on both species-specific interactions and landscape structure (e.g., riparian habitat and length of spawning reach) (Harding et al., 2019). Besides their well-documented effects during spawning seasons, inputs of necromass can have residual effects across seasons, with spawning salmon biomass in the autumn predictive of bird density and diversity in estuaries the following summer (Field and Reynolds, 2011). Findings such as these illustrate the important and complex roles necromass and phenology-based mortality play in aquatic ecosystems.

Although few aquatic insects have evolved to feed directly on carrion, the influx of nutrients from salmon carcasses, and resulting increases in primary production, can increase aquatic insect density by 8–25 times in artificial and natural streams where carcasses are present (Wipfli et al., 1998; Fenoglio et al., 2014). Isotopic studies have shown that salmon-derived carbon and nitrogen is incorporated into both primary producers and invertebrate feeding groups that consume microbes (e.g., filterers and grazers) (Bilby et al., 1996; Johnston et al., 1997; Guyette et al., 2014). While salmon remains the best studied example of phenology-based mortality (Schindler et al., 2003), they are by no means the only group of aquatic animals with programed mortality that leads to cascading effects in aquatic ecosystems. In addition to bony fish and invertebrates (see below), carcasses of other aquatic organisms, such as the sea lamprey (Petromyzon marinus), represent important nutrient sources, connect marine and freshwater systems, and can stimulate primary productivity (Weaver et al., 2018). While catadromous organisms which migrate to the ocean to spawn, [e.g., eels (Anguilla spp.)], also link marine and freshwater systems; how their behavior impacts marine ecosystems remains largely unknown. However, their spawning and subsequent death likely introduces considerable nutrients into otherwise oligotrophic environments where spawning occurs (e.g., Sargasso Sea).

### STOCHASTIC AND EPISODIC MASS MORTALITY AND DECOMPOSITION

### Vertebrate Carrion Mass Mortalities

One of the most striking examples of vertebrate mortality altering aquatic ecosystem comes in the form of episodic mass death, and the resulting carcasses that undergo decomposition, due to stochastic factors not easily predicted in nature (Fey et al., 2015). Mass fish kills are one of the most visible forms of this kind of mass mortality, where tens of thousands of fish may die within a short (e.g., hours to days) period of time, causing mass decomposition in the water and on banks of aquatic ecosystems (Ochumba, 1990; Thronson and Quigg, 2008). Fey et al. (2015) reported fish mass mortality events made up about 56% of all mass mortality events reported in scientific literature since the 1940s (N = 727). Episodic mass mortalities can be the result of stochastic changes in physical–chemical conditions (Cooper, 1993), toxic algal blooms (Hallegraeff, 1993), disease (Grizzle and Brunner, 2003), pollutants (Cooper, 1993), and other unknown factors. In many instances these conditions are the result of eutrophication over many years (Vollenweider, 1970; Harper, 1992; Nixon, 1995), or through punctuated high inputs of nutrients, like in the case of hippopotamus urine and feces (Subalusky et al., 2015) or mass drownings of wildebeest (Subalusky et al., 2017; Dutton et al., 2018). There are also other forms of vertebrate mass mortality that occur in aquatic ecosystems (Fey et al., 2015). For instance, annual mass drownings of wildebeest (Subalusky et al., 2017), aquatic reptiles (Rachowicz et al., 2006) and mammals that succumb to disease (Osterhaus et al., 1997; Kennedy, 1998) have both short- and long-term effects on ecosystem function. Fey et al. (2015) also provide an excellent quantitative assessment of mass mortality events worldwide, showing an increase in their occurrence for mammals, birds, amphibians, reptiles, fish, and marine invertebrates since 1940.

### Invertebrate Carrion Mass Mortalities

Much like cicada emergences that have been quantified to have significant impacts on terrestrial food webs and ecosystems (Yang, 2004), mayflies (Ephemeroptera), midges (Gratton et al., 2008), salmonflies (Plecoptera) (Walters et al., 2018; Wesner et al., 2019), and other aquatic insects (Baxter et al., 2005) emerge in masses as adults to mate and die, with their carcasses falling back to the aquatic habitat or into the adjacent riparian zones and inland landscapes (Gergs et al., 2014). When these mass emergences cross habitat or ecosystem boundaries (e.g., from a stream or lake onto the shoreline), such cross-ecosystem resource subsidies (Polis et al., 1997) can represent significant nutrient and energy pulses (Polis, 1994; Polis and Hurd, 1996). For example, Wesner et al. (2019) reported that for several stream sites with massive emergences of salmonflies (Pteronarcys californica), the resulting insect carrion deposition on the adjacent shore over only a single week was equal to or greater than annual atmospheric nitrogen and phosphorus deposition and the annual secondary production of all terrestrial insects from that watershed. Such contributions to the detrital pools of adjacent ecosystems have not been well studied, especially compared to the living emerged insects that are consumed by predators. Indeed, more studies are needed to better quantify the contributions of necromass originating from aquatic ecosystems and acting as resource subsidies to the decomposition budgets of adjacent ecosystems.

Mass mortalities of aquatic invertebrates, beyond what was discussed above with parasite-mediated terrestrial insect drownings, also affect in-stream and riparian communities and ecosystem properties. While not as well documented as fish subsidies, invertebrate mass mortalities have significant and sometimes long lasting effects on aquatic ecosystems since many of the invertebrate species, like mussels, have important functional roles (Vaughn, 2018), but also can contribute mass subsidies of highly recalcitrant structures, such as shells of molluscs (McDowell et al., 2017; DuBose et al., 2019). Beyond the pulsed effects of nutrient release related to rapid soft tissue decomposition of bivalves, Wenger et al. (2019) estimated that mussel shells from mortality events may have once provided about 1% of total phosphorus load in streams and rivers of the southeastern United States.

McDowell and Sausa (2019) reviewed the effects of mass mortality effects of the invasive bivalves Corbicula sp., the zebra mussel Dreissena polymorpha, the golden mussel Limnoperna fortunei, and the Chinese pond mussel Sinanodonta woodiana. Low and high water temperatures and water levels were the leading causes of mass mortality of both invasive and native species. There were short-term (i.e., over days) nutrient pulse releases associated with en masse soft tissue decomposition and longer-term microhabitat effects in the form of remaining shells both as part of the benthic substrata but also on stream and river banks. For instance, a mass mortality event of about 100 million C. fluminea contributed an estimated 751 kg of carbon, 180 kg of nitrogen, and 45 kg of phosphorus to a stream over the course of several days (McDowell et al., 2017). The dead and dying bivalve carcasses may also serve as food resources for local scavengers like fish, invertebrates, and birds (Mouthon and Daufresne, 2006; McDowell and Sausa, 2019). Mass mortality of C. fluminea left on stream banks attract a wide diversity of terrestrial invertebrates (Novais et al., 2015) and below ground nutrients (Novais et al., 2017), suggesting that bivalves washed onto banks during floods can become pulsed resources subsidies for adjacent habitats.

Mass mortality events also occur throughout the world in marine and estuarine habitats, and not surprisingly changing climate patterns and warming temperatures are increasing their magnitude and frequency (Coma et al., 2009). While not always independent of increases in water temperature, diseases also contribute to marine invertebrate mass mortality (Harvell et al., 1999). Not matter what the cause, marine invertebrate mass mortality has significant effects on coral reef and intertidal ecosystems by affecting both hard and soft corals, benthic burrowers and filterers with cascading effects on upper trophic level consumers (Knowlton, 2004). These massive death events are often extensive. For instance, in 2016, one of the three pantropical mass coral bleaching events occurred in the Great Barrier Reef of Australia as a result of a significant marine heat wave,

where 90% of the surveyed reefs suffered mortality (Hughes et al., 2017). The ecosystem consequences of this event included significant community restructuring, functional changes, and widespread declines in consumers resulting directly from the water temperature increases, but also in relation to coral loss (Stuart-Smith et al., 2018). Increased frequency and magnitudes of such global weather events will have widespread ecological impacts on all ecosystems, with mass mortalities contributing to many of the most negative effects of a changing climate.

#### SUMMARY

Death and decomposition occur in all ecosystems, but the extent and magnitude of the resulting necromass varies in space and time. Some forms of necromass come from within the system, as part of the life cycles and life histories of the resident organisms (e.g., planktonic snow in a lake or ocean or insects in a stream) or during mass deaths resulting from changing and intolerable habitat conditions (e.g., fish kills and coral reef bleaching). Other forms of necromass subsidies come from outside of the system, sometimes as migratory fish (e.g., salmon) or parasitized terrestrial insects that have been behaviorally hijacked to enter aquatic habitats where they drown. In all cases, the once living biomass becomes an often significant and functionally important component of the detrital pool. This component of heterotrophically derived necromass has historically been difficult to quantify due to the rapid turnover of such labile resources, and has arguably been overlooked or underappreciated in many ecosystem level energy and nutrient budgets. Recent work has provided conceptual models and methods for improving the ability to identify, quantify, and better study how carrion resource subsidies affect aquatic ecosystems, ranging from small, headwater streams to saltmarshes and mangroves, to the deep oceanic abyss and enormous stretches of coastline habitats around the world. Furthermore, while studies of megafauna carrion (e.g., whales and whale sharks) have demonstrated significant impacts to deep ocean habitats, much less is understood for

#### REFERENCES

other groups of animals. Quantifying the energy, nutrient, and foodweb effects of autochthonous and allochthonous carrion resulting from non-consumptive morality, phenologybased mortality and stochastic and episodic mortality events will allow broader assessments of all necromass contributions in aquatic ecosystems. With advances in geographical (e.g., drones and satellite imagery), genomic (e.g., next-generation sequencing), and other forms of technology (e.g., cell phone cameras and associated software) it will become increasingly feasible to quantify baseline levels of carrion to more fully evaluate how this resource subsidy affects ecosystem energy and nutrient dynamics.

#### AUTHOR CONTRIBUTIONS

MB, JR, and GL performed the literature search, and writing and editing of this review.

#### FUNDING

MB was supported by the College of Agriculture and Natural Resources and College of Osteopathic Medicine of Michigan State University. GL was supported by the Department of Biological Sciences and the Environmental Change Initiative at the University of Notre Dame.

#### ACKNOWLEDGMENTS

We would like to thank P. Barton, J. Tomberlin, H. Jordan, and A. Tarone for thoughtful discussions that have led to this manuscript, and J. Pechal also for reviewing an earlier version. This manuscript was the ultimate product of a special session at the 2018 Annual Meeting of the Society for Freshwater Science, hosted in Detroit, MI (https://freshwater-science.org/ annual-meeting) by MB and GL, where many other scientists inspired our writing through their presentations and manuscript contributions to this special issue.

decomposition ecology of autotrophically and heterotrophically derived organic matter. Ecol. Monogr. 89:e01331. doi: 10.1002/ecm.1331

Front. Ecol. Environ. 1, 31–37. doi: 10.1890/1540-92952003001[0031:PSATEO] 2.0.CO;2

**Conflict of Interest:** The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Copyright © 2020 Benbow, Receveur and Lamberti. This is an open-access article distributed under the terms of the Creative Commons Attribution License (CC BY). The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

# Quantitative Food Webs Indicate Modest Increases in the Transfer of Allochthonous and Autochthonous C to Macroinvertebrates Following a Large Wood Addition to a Temperate Headwater Stream

#### Edited by:

Sally A. Entrekin1,2

Gary A. Lamberti1

\*, Emma J. Rosi3

Carla Atkinson, The University of Alabama, United States

#### Reviewed by:

Arthur Benke, The University of Alabama, United States Caleb Robbins, The University of Kansas, United States

\*Correspondence:

Sally A. Entrekin<EMAIL_ADDRESS><EMAIL_ADDRESS>
#### Specialty section:

This article was submitted to Population and Evolutionary Dynamics, a section of the journal Frontiers in Ecology and Evolution

Received: 05 September 2019 Accepted: 08 April 2020 Published: 06 May 2020

#### Citation:

Entrekin SA, Rosi EJ, Tank JL, Hoellein TJ and Lamberti GA (2020) Quantitative Food Webs Indicate Modest Increases in the Transfer of Allochthonous and Autochthonous C to Macroinvertebrates Following a Large Wood Addition to a Temperate Headwater Stream. Front. Ecol. Evol. 8:114. doi: 10.3389/fevo.2020.00114 1 Department of Biological Sciences, University of Notre Dame, Notre Dame, IN, United States, 2 Department of Entomology, Virginia Tech, Blacksburg, VA, United States, 3 Cary Institute of Ecosystem Studies, Millbrook, NY, United States, 4 Department of Biology, Loyola University, Chicago, Chicago, IL, United States

, Jennifer L. Tank1

, Timothy J. Hoellein1,4 and

Headwaters suffer from reduced leaf and wood inputs and retention capacity from historical land actions like watershed logging and agriculture. When in-stream wood is reduced, stream retention capacity declines and subsequent changes in streamwater flow-paths and patterns of deposition alter decomposition and primary production that influence secondary invertebrate production via modified habitat and resources. Wood additions are commonly used as stream restoration tools for habitat improvements that can restore or strengthen food web connections; however, changes in carbon (C) flow through food webs are rarely measured because of time and expense. We quantified allochthonous and autochthonous C flow through aquatic macroinvertebrate communities 1 year before and 2 years after an experimental addition of large wood, compared to macroinvertebrates in an upstream control, in a temperate headwater stream. We predicted wood additions increase macroinvertebrate consumption and assimilation of allochthonous and autochthonous C through retention of leaves and altered flow-paths that expose more gravel and cobble for periphyton colonization. Macroinvertebrate allochthonous C assimilation tended to increase in years with greater organic matter retention and autochthonous C increased with more exposed gravel and cobble across seasons and between reaches. While the effect of wood addition on C flow through the macroinvertebrate community was minimal, it increased by ∼20% relative to the control from an increase in production and C assimilation of common mayfly and caddisfly scrapers, Baetis and Glossossoma. Because the amount of organic matter retained and coarse substrate exposed corresponded with C form and amount consumed, restoration of large wood has the potential to increase organic matter C trophic transfer.

Keywords: headwaters, large woody debris, restoration, stable isotopes, organic matter

### INTRODUCTION

fevo-08-00114 May 4, 2020 Time: 17:33 # 2

Freshwater ecosystems cover only a small fraction of Earth's surface but receive at least 1.9 Pg carbon yr−1 from terrestrial ecosystems. At least 50% of these carbon (C) inputs are stored in or emitted from freshwaters (Cole et al., 2007). Consequently, cross-ecosystem terrestrialaquatic exchanges are essential for understanding global C cycling (Cole et al., 2007). Stream food webs are strongly influenced by leaf-litter and wood from adjacent riparian areas. Changes in riparian detrital quality and quantity dictate aquatic biological community structure, organismal growth, and organismal lifecycle completion rates, described as organismal performance (Wallace et al., 1997b). Collectively, this community performance governs aquatic ecosystem functions like secondary production and decomposition (Webster et al., 1997). In turn, in-stream secondary production and decomposition govern the capacity of small streams to transfer and transport C that subsidizes downstream and riparian communities.

Human actions in forested watersheds can change the timing or amount of allochthonous and autochthonous C that affect the organisms available to consume and transfer C (Cummins et al., 1989). For example, greater retention of allochthonous material in streams stimulate growth of fungi and bacteria, and organic matter consumption by macroinvertebrates (Richardson, 1991; Negishi and Richardson, 2003; Tiegs et al., 2008), increasing the relative amount of allochthonous versus autochthonous C transferred through the stream food web (Rosemond et al., 1993). However, increases in water velocity can increase exposure of large substrates subsequently covered in sand and stimulate periphyton growth (Kail, 2002), increasing the relative contribution of autochthonous C to secondary consumers like macroinvertebrates (McNeely et al., 2007). Large wood results in both organic matter retention and substrate sorting via modified flow paths. Therefore, adding wood to streams can change available habitat and food resources for aquatic biota. The concurrent increase in autochthonous and allochthonous C in low-production, forested streams could lead to an overall increase in C contributions to secondary macroinvertebrate production, resulting in greater trophic transfer and more trophic linkages.

Attempts to restore headwater streams through a bottomup organic matter addition (e.g., addition of large wood) provide stability and greater resource availability that could also alter the relative terrestrial- and instream-derived C eaten and assimilated by aquatic heterotrophs (e.g., Rosi-Marshall and Wallace, 2002). Forested headwater streams have historically been considered to be strongly influenced by the volume and timing of terrestrially derived allochthonous material delivered to the stream (Polis and Strong, 1996; Wallace et al., 1997b; Moore et al., 2004). However, algae can also be seasonally important in many temperate streams (Finlay, 2001; Hall et al., 2001). The alternating increase of algae in winter when the canopy is more open and leaf litter inputs retained in autumn may act to stabilize the food web when resources may otherwise be scarce (Power et al., 1988; Polis and Strong, 1996; Moore et al., 2004; Guo et al., 2016). We are unaware of any studies that have quantified allochthonous C (i.e., terrestrially derived) versus autochthonous C (algal-derived) following instream wood addition.

Ecosystem processes that restorations aim to affect include rates of nutrient uptake (Sudduth et al., 2011), organic matter retention, and decomposition (Lepori et al., 2005; Frainer et al., 2017). These functional metrics are especially useful in restoration studies because they reveal how changes in physical structure may influence the rate, and pathway of energy or elements flowing through an ecosystem. Stable isotopes can be used to identify the source and amount of C assimilated by a target community and are increasingly used in restorations (for example, Fry, 2002; Kennedy et al., 2005; Lepori et al., 2006). For example, Lepori et al. (2006) added boulders to several streams in Sweden that increased detrital retention, but did not result in consumers δ 13C more similar to the retained detritus.

Still, stable isotopes may identify food source assimilated and can be used in combination with measures of community structure and secondary production to develop quantitative food webs (e.g., Rosi-Marshall and Wallace, 2002). Quantitative food webs can reveal consumer-level controls on ecological processes and illustrate changes in trophic structure (Rosi-Marshall and Wallace, 2002), community assemblage, and nutrient flow (Cross et al., 2007)following ecosystem restoration. Tracking pathways of energy flow through the food web integrates changes in food resource assimilation, community structure, survival, and production (Benke and Wallace, 2011). Carbon flow measurements are commonly used to test ecological theory (e.g., Cross et al., 2007), but have not been used to assess restorations.

We quantified allochthonous and autochthonous C assimilation in macroinvertebrates before and after experimental wood addition in a Michigan headwater stream using natural abundance of C isotopes for macroinvertebrates and their food resources. Then we combined estimates of C assimilation with secondary production (Entrekin et al., 2009) to calculate trophic basis of production (Benke and Wallace, 1997). Finally, by using assumed assimilation efficiencies, we were able to back-calculate the amount of allochthonous and autochthonous C consumed (expressed as a rate) and compare that C flow with the amount available. Our previous work in this stream showed that total macroinvertebrate secondary production was low before wood addition, but increased by ∼25% 2 years after wood addition, resulting in a statistically significant increase in invertebrate biomass and greater secondary invertebrate production (Entrekin et al., 2009). Here, we sought to quantify how assimilation, trophic basis of production and C allochthonous and autochthonous C flowing through the macroinvertebrate food web changed after wood was added. We predicted wood addition would increase consumption and assimilation of allochthonous and autochthonous C by macroinvertebrates through an increased retention of leaf litter, and the exposure of sand-covered large inorganic substrates for periphyton colonization.

## STUDY SITE

fevo-08-00114 May 4, 2020 Time: 17:33 # 3

State Creek is a 1st-order stream draining 3.9 km2 in the Ottawa National Forest in the Ontonagon River basin in the Upper Peninsula of Michigan, United States (46◦ 280N, 89◦ 1 0W). We began sampling monthly in May 2003, 1 year before wood addition (Y0) and continued for 1 (Y1) and 2 (Y2) years after wood addition in both the wood-added (treatment) and the upstream control reach. Treatment and control reaches were separated by a 50 m distance to promote some independence between sites, while reducing differences associated with longitudinal changes. We measured no difference in the amount of in-stream wood between the control and treatment reaches prior to the wood addition (Entrekin et al., 2007). The stream section we studied had 90% canopy cover during spring and summer and 75% canopy cover in autumn and winter, an average bank-full width of 2.4 m in the control reach and 2.5 meters in the treatment reach, an average water depth at base flow of 13 cm in the control reach and 12 cm in the treatment reach, and an average discharge in the control reach of 64 and 67 L s−1 in the treatment reach (for more details see Entrekin et al., 2007; Hoellein et al., 2009). The stream flows through a managed, second-growth forest with intact, but young, riparian vegetation in a catchment of 95% deciduous forest that was last logged in 1967. Riparian trees include Populus tremuloides Michx (trembling aspen), Acer rubrum L. (red maple), Acer saccharum Marsh. (sugar maple), Betula papyrifera Marsh. (paper birch), Tsuga canadensis L. (hemlock), Pinus alba L. (white pine), with a thick understory of Alnus serrulata Alt. (tag alder). The study stream had low in-stream large wood density (13 pieces/100 m stream length) and low storage of coarse benthic organic matter [annual average of 146 ± 70 (SE) g AFDM m−2 ] before wood addition (Cordova et al., 2007; Entrekin et al., 2007) from a history of region-wide logging and shale mining (Webster et al., 2008).

In May 2004, we added 25 logs (each 2.5 m long × 0.5-m diameter) of big tooth aspen P. grandidentata Michx., purchased from a nearby tree farm, haphazardly to a 100-m stream reach, while maintaining a 100 m upstream control reach. After 2 years in the stream, 15 of the 25 logs moved. Most of the added logs moved less than two meters with one moving the farthest at 18 m (G. Lamberti unpublished data). None of the added wood moved out of the study reaches. We did measure an increase in the amount of organic matter retained and more exposed cobble from a localized increase in water velocity caused by the added logs (Entrekin et al., 2008).

### MATERIALS AND METHODS

#### Habitat Characteristics

Standing crops of coarse (CBOM) and fine (FBOM) benthic organic matter were measured from five 804-cm2 benthic cores sampled in each reach on each sampling date; CBOM was separated from FBOM using a 1-mm sieve. After CBOM was removed from the corer, a FBOM slurry was made by stirring the sediment in the core and then subsampled using a 160 mL specimen container. Subsamples were stored on ice until processing. In the laboratory, samples for CBOM were dried at 60◦C, sorted by organic matter type (leaves, moss, and wood), and weighed. A subsample of each organic matter type was then combusted at 550◦C, and reweighed to determine ash-free dry mass (AFDM; Benfield 2006). For FBOM, subsamples were filtered onto glass fiber filters (GF/F), dried at 60◦C, weighed, combusted at 550◦C, and reweighed for AFDM.

On each sampling date, we also measured discharge and velocity from the dilution of a conservative tracer during concurrent measurements of whole-stream nutrient uptake rates (Hoellein et al., 2007). Water temperature was recorded at the bottom of each stream reach hourly from May 2003–May 2006 using HOBO R data loggers (Onset Computer Corporation, Bourne, MA, United States). We also surveyed benthic habitat using transects spaced every 5 m (perpendicular to flow) in both reaches in May and August of 2003 (before wood addition) and in May, July, and November 2004–2006 (after wood addition). Sediments were categorized using the Wentworth scale (Minshall and Rugenski, 2006). Inorganic sediments were classified as boulders, gravel and cobble, and sand, while organic substrates were moss, CBOM (>1 mm), silt, small wood (<10 cm), and large wood (>10 cm). Measurements were recorded every 20 cm across the channel. We calculated percent cover for each substrate at each transect-scale, and as the mean of each category across all transects combined (i.e., reach-scale).

#### Measuring δ 13C Natural Abundance Signature

We measured δ 13C values for the most productive taxa that collectively represented ∼90% of the total macroinvertebrate community production (Entrekin et al., 2009). Macroinvertebrate abundance, community composition, and secondary production were measured from monthly (none collected in January) Hess (32 cm diameter, 250 µm-mesh) samples (5 per stream reach) in the treatment and control reaches for the 5-year study period. For secondary production, size-frequency histograms were developed for each taxon and corrected using cohort production intervals. For rare taxa, we used either production to biomass ratios we developed or published values (Entrekin et al., 2009). We then used a sub-set of those individuals for stable isotope analysis. Samples were preserved in 6–8% formalin, which we note may result in a 1–1.65h systematic depletion of δ 13C across taxa (Sarakinos et al., 2002; Bicknell et al., 2011). However, a comparison of δ 13C differences between frozen and formalin-preserved samples for four dominant taxa in one season showed no consistent change (S. Entrekin, unpublished data). Despite some inorganic C, we did not acidify samples because inorganic C was low and acidifying can lead to fractionation (Schlacher and Connolly, 2014). To measure C stable isotopes of macroinvertebrates, we selected late instar taxa from March (late winter), May (spring), June or July (summer), and November (autumn) for the three study years to represent possible seasonal changes in taxa. Thus, our macroinvertebrate sampling incorporated seasonal variation in macroinvertebrate diet as well as any potential changes due to the wood addition.

For each stable isotope measurement, the number of individuals representing a single taxon varied based on an individual's mass (i.e., 2–20 individuals). When possible, we used individuals from at least three different replicate Hess cores that were taken haphazardly along each 100-m reach to incorporate reachscale variation and expressed as averages with standard error (see **Supplementary Appendix 1**). Finally, we measured the δ 13C signature of macroinvertebrate food resources including conditioned leaves (i.e., leaf litter colonized by bacteria and fungi) and algae from grab samples in the selected months. Conditioned leaves were collected along each stream reach from State Creek in spring, summer, autumn, and winter and frozen until analysis. Periphyton samples were also collected seasonally by scraping multiple rocks throughout the stream reach.

Macroinvertebrate taxa and food resources were dried at 60◦C, ground to a fine powder, and analyzed on a Finnigan Delta Plus Stable Isotope Mass Spectrometer. Stable isotope values were expressed in δ notation as the difference in parts per thousand (h) from a standard (PeeDee Belemnite), using the calculation: δ 13C = (Rsample/Rstandard−1) ∗ 1000; where R is 13C/12C. Our attempts at silica separation (Hamilton et al., 2005) were unsuccessful in isolating low amounts of algae (mostly diatoms) from periphyton biofilm scrapings. Therefore, we used the δ 13C value of a known grazer, the caddisfly Glossosoma sp., as the presumed proxy for periphyton, as has been done in other studies (e.g., Finlay, 2001). Glossosoma gut contents were examined each season to verify the presence of algae; however, small amounts of allochthonous C may have been assimilated that would result in an underestimate of the contribution of autochthonous C (to the invertebrate community diet).

Our first goal was to partition allochthonous (i.e., leaf litter) from autochthonous (i.e., in-stream algae) C assimilation by macroinvertebrates as indicated by δ 13C signatures using a twosource mixing model rather than using gut content analysis. Therefore, no other food resources were used in the mixing model because leaf litter and epilithic algae (or periphyton) represented the end members (allochthonous versus autochthonous food resources) in this food web. We used the model IsoSource and IsoError from the Environmental Protection Agency1 , which has the advantage of accounting for replicate sample variation (Phillips and Gregg, 2001). The δ 13C average of conditioned leaf litter was −29.98 ± 0.01h, and δ 13C of algal signature of Glossosoma sp. was 36.94 ± 0.2h (upper confidence interval = 1 and lower confidence interval = 0.97) across our sample period, indicating separation of the two C sources. δ 13C signatures were not corrected for C fractionation rates as they are likely low and unpredictable (McCutchan et al., 2003).

#### Relative Allochthonous and Autochthonous C Assimilation and Their Contribution to Production

Then, we used a modified method to quantify the trophic basis of production described by Benke and Wallace (1997, 2011) where the proportional contribution of allochthonous and autochthonous C to assimilation for each taxon, as indicated by stable isotopes, was multiplied by the total production of that taxon to determine the contribution to secondary invertebrate production or the trophic basis of production. The contribution of allochthonous and autochthonous C to production was then calculated for each dominant macroinvertebrate taxa and summarized for each functional feeding group (FFG). The advantage of this method was that assumptions about assimilation efficiency were not needed because natural abundance of isotopes reflect assimilation. However, we were not able to estimate total amount of resources consumed. Therefore, to quantify how much autochthonous and allochthonous food was consumed (i.e., in units of AFDM per m−2 time−1 ), we divided secondary production by published assimilation efficiencies for detritus and algae and then multiplied that by published net production efficiency (Benke and Wallace, 1997; Hall and Meyer, 1998; Rosi-Marshall and Wallace, 2002) to get autochthonous and allochthonous C consumption (Benke and Wallace, 1997):

Algal or detrital C consumption = (secondary production/assimilation efficiency) ∗net production efficiency where secondary production is expressed as mg dry mass per m−2 yr−1 , assimilation efficiency is a proportion (detritus = 0.1 or algae = 0.3), and net production efficiency is a proportion (NPE = 0.5).

#### Statistical Analyses

We used a Before-After-Control-Impact Analysis of Variance (BACI-ANOVA) (Stewart-Oaten et al., 1986; Underwood, 1992) with a Tukey post hoc when significance at α = 0.05 was met to test for changes in δ 13C and the contribution of allochthonous versus autochthonous C for the macroinvertebrate community, functional feeding groups, and for individual taxa. Pearson product moment correlations were used to explore the relationship among δ 13C values and environmental parameters that may contribute to variation across stream reaches through time (e.g., organic matter standing crops and % substrate cover). All data were tested for conformance to homogeneity of variance using Levene's test and assumptions of ANOVA using Kolmogorov-Smirnov tests, and transformed when assumptions of normality were not met. All analyses were performed with SAS software (v. 8.02, SAS Institute Inc., Cary, NC, United States).

The amount of C consumed, assimilated and supporting secondary production (i.e., flux or flow) were not replicated measurements; therefore, we did not conduct statistics, but we present and describe the effect size as the difference between the treatment and control reaches before and after wood addition.

#### RESULTS

#### Changes in δ 13C Signatures After Wood Addition

δ 13C ranged from ∼−25h to −37h across all macroinvertebrate taxa (**Supplementary Appendix 1**). We were surprised that overall average macroinvertebrate δ 13C remained mostly unchanged following wood addition (**Figure 1**). Before wood

1http://www.epa.gov/wed/pages/models/stableIsotopes/isotopes/isoerror1\_04. htm

addition, the mean (±1 standard error) δ 13C signature across all macroinvertebrate taxa differed by ∼1h between the control and treatment reaches (**Figure 1** and **Table 1**). The pattern remained after wood addition, and there was no difference in community-level δ 13C between the treatment and control reach in Y1 (F3,12 = 0.42, P = 0.74) or Y2 (F3,12 = 0.23, P = 0.87) (**Table 1**). Similarly, there was no difference in functional feeding group and taxa-specific δ 13C between reaches after wood addition in Y1 or Y2 (**Table 1**).

#### Seasonal Changes in δ 13C Signatures

Macroinvertebrate δ 13C signatures reflected seasonal variation in resources as expected in a temperate headwater stream with signatures closer to leaf-litter in autumn (**Supplementary Appendix 1** and **Figure 1B**) and farther from leaf-litter in March (**Figure 1C**) when canopy is more open. Community mean δ 13C values were different among seasons (F3,8 = 13.54, P < 0.001) with greater δ 13C value of −31.79h (i.e., closer to leaf litter signature) in autumn (November) and a lower ∼−33.5h δ 13C in winter, spring, and summer (i.e., farther from leaf litter signature) (**Figure 1**, Tukey's test P < 0.001).

The δ 13C signatures of many individual taxa responded to wood addition differently depending upon taxonomic identity and season, although not necessarily in a predictable manner (**Supplementary Appendix 1** and **Table 1**). For example, Tipula, an obligate shredder, tended to have δ 13C closer to leaf litter in the treatment reach relative to the control in Y2 compared to their difference in Y0. A dominant filterer, Simulium, had lower δ 13C in the treatment reach in May of Y2 compared to the difference between the control and treatment reach in Y0 (**Supplementary Appendix 1**, F3,4 = 77.77, p < 0.001). In addition, the predatory stonefly, Isogenoides, had greater δ 13C in the treatment reach in autumn of Y1 compared to the difference between the control and treatment reach in Y0 (**Table 1**, F3,5 = 11.7, p = 0.01). Baetis, among the most productive facultative scraper and gatherers, had greater δ 13C in the treatment reach in Y2 winter compared to the difference between the control and treatment reach in Y0 (**Supplementary Appendix 1**, F3,7 = 18.6, p = 0.001).

#### Variation in Macroinvertebrate Functional Feeding Group C Assimilation Related to Substrate Composition and Organic Matter Standing Stocks

We correlated allochthonous C assimilation of average community and functional groups with substrate type to identify environmental factors influencing variation across stream reaches and season (**Figure 2**). The overall % allochthonous C assimilation at the community level was not explained

#### TABLE 1 | Mean annual δ 13C values for each taxon.

fevo-08-00114 May 4, 2020 Time: 17:33 # 6

\*Indicates a significant change in mean macroinvertebrate δ 13C value after wood addition. D is the difference in treatment (T) and control (C). B is before (Y0) and A is after (Y1 and Y2) wood addition. Functional groups are in bold.

by any one variable, although variation between stream reaches was related to the proportion of sand making up the stream bottom (**Figure 2A**). The relative amount of allochthonous C assimilated by macroinvertebrate FFGs across season and reaches was mostly correlated to differences in available coarse substrate and organic matter standing crop. The overall proportion of allochthonous C assimilation by shredders, filterers, and gatherers was positively related to the mean annual coarse organic matter standing crop (**Figures 2B,C,E**). In contrast, scrapers, that feed mostly on inorganic substrates, ranged from less than 1 to 15% allochthonous C assimilation. The proportion of terrestrially derived C assimilated by predators closely tracked gatherer assimilation of allochthonous C (**Figure 2F**).

## All Functional Feeding Groups Consumed Allochthonous and Autochthonous C Sources, but

FFGs Consumed Different Amounts of the C Sources Allochthonous and autochthonous-based C contributed roughly equal parts to macroinvertebrate production in the control and treatment reaches, but with little change following wood addition. Annual contribution of allochthonous- C ranged from 35 to 44% in the control reach and 41–56% in the treatment reach over the 3-year study (**Figure 3A**). The origin of C assimilated by FFGs differed among groups as expected. Autochthonousbased C contributed the most to scraper production, ranging from 70 to 100%. However, allochthonous-based C contribution decreased in scrapers in Y1 relative to the difference between the control and treatment in Y0 (**Figure 3B**). For shredders,

contribution of allochthonous C ranged from 50 to 90%, varying among years, reach, and taxa (**Figure 3C** and **Supplementary Appendix 1**). Contribution of allochthonous C to gatherer production increased in Y1 in the treatment reach relative to the difference between the control and treatment in Y0, ranging from 58 to 66% in the control reach and 72 to 88% in the treatment reach (**Figure 3D**). Allochthonous C contribution to filterer production increased in both reaches in Y1 and declined back to pre-wood addition values in Y2 (60–90%; **Figure 3E**). Allochthonous C contribution to predator production ranged from 62 to 77% and changed very little following the addition of wood (**Figure 3F**).

### Actual Allochthonous-Based and Autochthonous-Based C Contributing to Macroinvertebrate Production

In the control reach, allochthonous and autochthonous production declined each year of the study. In contrast, allochthonous and autochthonous -based production in the treatment reach was identical in Y0 and Y2 but lower in Y1 (**Table 2** and **Figure 4A**). All FFG secondary production declined in Y1, while scraper and shredder production increased beyond Y0 values. Autochthonous-based scraper production was lowest in the treatment reach in Y1 and greatest in the treatment reach in Y2 (**Figure 4B**). Changes in scraper production in Y2 of the treatment reach were driven by an increase in mayfly Baetis production and C assimilation changed from mostly autochthonous to more allochthonous (**Table 2**). Allochthonous based shredder production was also lowest in Y1 and greatest in the treatment reach in Y2 from an increase in Tipula production (**Figure 4C**). Other FFGs did not show evidence of a treatment effect. Gatherer production was consistent across years in the control reach, but declined in the treatment reach, while filterer production declined across years in both reaches (**Figure 4E**). Finally, overall predator production and the ratio of allochthonous and autochthonous-based C contributions to production changed little across dates and reaches (**Figure 4F**).

### Consumption of Allochthonous and Autochthonous- C

We calculated the amount of allochthonous and autochthonous C consumed by macroinvertebrates from published assimilation efficiencies to document changes in the amount of each resource flowing through the macroinvertebrate food web. Because leaf litter has a lower assimilation efficiency (0.1) than algae (0.3), macroinvertebrates must consume more allochthonous -based C to support their energetic demands (Benke and Wallace, 1997). Therefore, autochthonous-based C contributed a larger proportion to production, even though autochthonous consumption was less than allochthonous (**Figure 5A**). All functional feeding groups consumed both C sources. However, scrapers consumed relatively more autochthonous (**Figure 5B**) and filterers (**Figure 5C**) consumed more allochthonous C. Shredders (**Figure 5D**) and gatherers (**Figure 5E**) consumed mostly allochthonous C and contributed more to overall energy

flow by consuming more material to support relatively greater secondary production (**Figure 5**).

#### Energy Flows From Autochthonous and Allochthonous-Based C

In the control reach, the dominant energy flows remained fairly consistent among Y0, Y1, and Y2 (**Table 3** and **Figure 6**). autochthonous C consumed by scrapers remained fairly evenly distributed between Baetis and Glossosoma in each year, with a slight increase to Baetis in Y2 (**Figure 6**). Allochthonous production also remained constant flowing mostly through the filtering black flies, Simulium, and to a lesser extent a netspinning caddisfly, Parapsyche, with no changes from Y0 to Y2. In contrast, energy flow in the treatment reach showed much greater variation across years (**Figure 6**). Scraper basal C flow in Y0 was split between Baetis and Neophylax in the treatment reach and was predominantly autochthonous C sources. In Y1 and Y2 in the treatment reach, allochthonous C contributed the most to Baetis production in Y2 (**Figure 6**). For shredders, C flow through allochthonous C flow through Tipula was greater in Y2 than in Y0 or Y1.

TABLE 2 | Total allochthonous (alloch) or autochthonous (auto) carbon based production (mg AFDM m−2 y−1) of each taxon calculated as production multiplied by the proportion of alloch or auto carbon source assimilated (estimated fromδ13C values).

\*Indicates a significant change in mean macroinvertebrate δ13C value after wood addition. D is the difference in treatment (T) and control (C). B is before (Y0) and A. is after (Y1 and Y2) wood addition. Functional groups are in bold.

(Y1) and 2 years (Y2) after wood addition. (A) Is total macroinvertebrate production weighted for resource assimilation, (B) is scraper production weighted for resource assimilation, (C) is shredder production weighted for resource assimilation, (D) is gatherer production weighted for resource assimilation, (E) is filterer

production weighted for resource assimilation, and (F) is predator production weighted for resource assimilation.

DISCUSSION

#### Implications for Management and Restoration

Combining δ 13C signatures to get allochthonous and autochthonous C assimilation with secondary production provided a unique investigation into the macroinvertebrate trophic response to wood addition (**Figure 6**). The combined approach of C assimilation and macroinvertebrate secondary production indicated changes in the consumption and flow of allochthonous and autochthonous C through the macroinvertebrate food web. Here we demonstrated that the abundance of organic matter and locally exposed cobble and gravel, which supports periphyton, could drive increased allochthonous and autochthonous C flow to several macroinvertebrate functional feeding groups most notably greater autochthonous C contribution to scrapers and greater allochthonous C contribution to filterers shredders (**Figure 3**). Taxon-specific seasonal and annual shifts in C assimilation were documented (**Supplementary Appendix 1**). Still, we observed changes in total resource consumption from the coupled increase in organic matter standing crop and substrate sorting at a local scale (i.e., around the added logs) following wood addition (Entrekin et al., 2008). Unfortunately, we did not replicate this study, which is a commonly encountered challenge for reachscale quantitative food web approaches that measure energy flow (e.g., Cross et al., 2007). Despite the limits for generalizations, our results support methods for future evaluations of restoration (see also Vander Zanden et al., 2006) by stressing the need for trophic-based analysis to assess restoration success or failure in terms of resource availability and energy flow. The combined structural and functional approach here points to physical changes from the restoration that increased allochthonous and autochthonous C flow (**Figure 5**) and increased food web

TABLE3|Totalamountofallocthonous(alloch)andautochthonous (auto)foodsourcesconsumedbyeachtaxon(mgAFDMm−2y−1).

\*Indicates a significant change in mean macroinvertebrate δ13C value after wood addition. D is the difference in treatment (T) and control (C). B is before (Y0) and A. is after (Y1 and Y2) wood addition. Functional groups are in bold.

connections predicted from large wood additions in temperate streams (**Figure 6**).

Modest changes in C flow following wood additions were from a combination of greater assimilation of allochthonous and autochthonous C by a few macroinvertebrate taxa. Baetis mayflies, a multivoltine taxa, responded with greater secondary production that was fueled by increasing amounts of allochthonous C. Glossosoma caddisflies also showed an increase in production fueled by autochthonous C (**Figure 6**). Allochthonous and autochthonous resources were nearly equally important in this forested temperate stream. Typically, allochthonous resources are most important and organic matter standing crop predicts macroinvertebrate secondary production (Wallace et al., 1987, 2015); however, in the main channel of our stream, autochthony was as well documented. For example, 2 years after the wood additions, ∼40–50% of the macroinvertebrate community was supported by autochthonous C (**Figure 3**). In contrast to a manipulation of allochthonous C in an Appalachian stream (Wallace et al., 1997b), autochthonous C was a more important and substantial resource to the food web (**Figures 4**, **5**). The difference in our study compared to studies in the Appalachian stream is the absence of a dense understory of rhododendron and a mature forest (Greenwood and Rosemond, 2005; Wallace et al., 2015). In the Appalachian streams, rhododendron and a more mature forest canopy limited light for primary production (Greenwood and Rosemond, 2005). Light was not limiting during seasons of open canopy in our study stream and this was likely the primary reason for

autochthonous C.

greater autochthonous resource availability in this study stream compared to others conducted in headwater streams draining in-tact forests (Hoellein et al., 2007). Still, allochthonous C was the primary C source contributing to macroinvertebrate secondary production, particularly for shredders (∼50–80% contributed to secondary production), gatherers (∼60–90%), filterers (∼55–90%), and predators (∼60–80%) (**Figure 3**). The overall contribution of allochthonous C to secondary production ranged from 50 to 60% because scraper production was relatively high compared to the other functional groups (**Figure 4**).

#### Seasonal Effects of Added Wood on Macroinvertebrate δ 13C

While annual average macroinvertebrate isotopic signatures changed only a little through space and time, taxon-specific signatures did change seasonally (**Figure 1**). Taxa with seasonally variable δ 13C signatures may act as 'indicators', useful in evaluating ecosystem restorations aimed at changing or increasing basal resources [sensu 44]. In this study, Glossosoma, considered obligate consumers of algae, showed little variability in δ 13C across months, years, and stream reaches; therefore, their production rather than their δ 13C signatures may serve as a metric of restoration success (**Table 1** and **Supplementary Appendix 1**). In contrast, facultative feeders such as the shredding stonefly Nemoura sp., and Baetis spp., showed spatially and temporally variable δ 13C values, and so their average δ 13C values at a population level may be more indicative of seasonal changes in basal food resource availability. Using the natural abundance of C stable isotopes for assessing resource acquisition complements assessments based on community-level diversity and tolerance values (e.g., Gratton and Denno, 2006).

## Influence of Large Dead Wood on Macroinvertebrate C Flow

Given the forested nature of our study stream and the well-documented role of leaves and wood in governing macroinvertebrate secondary production, we hypothesized that most C in stream macroinvertebrates would bear an allochthonous signature (Wallace et al., 2015). Indeed, we predicted primary production would be limited by the deciduous canopy that was 75% closed in the winter and spring and 90% closed in the summer and autumn (Entrekin et al., 2008). Furthermore, a comparison of organic matter standing crop across studies and years show a strong positive relationship with macroinvertebrate secondary production suggesting reliance on allochthonous C as a primary food resource (e.g., Newbold et al., 1997; Webster et al., 1997; Hall et al., 2001; Chadwick and Huryn, 2007; Cross et al., 2007; Entrekin et al., 2007). However, these studies did not measure assimilation and algae can contribute more to production than often assumed from feeding mode analysis and correlations to resource availability (Finlay, 2001; Hall et al., 2001; Marcarelli et al., 2011). Therefore, we were surprised to find that the autochthonous C supported 47–59% of secondary macroinvertebrate production in both reaches (**Figure 3**).

The few studies that measured the contribution of primary production to macroinvertebrates in temperate headwater streams have based their conclusions on gut content analysis or tracer stable isotopes. For example, Mayer and Likens (1987) used gut contents to conclude that algae were more important than expected for caddisflies in forested headwater streams. More recently, allochthonous and autochthonous C resource contributions have been delineated for entire macroinvertebrate communities in a variety of biomes using stable isotopes with the consensus being algae were more important to the food web than expected (Guo et al., 2016; Brett et al., 2017; Neres-Lima et al., 2017). In fact, macroinvertebrates sampled from headwaters in the Salmon River, where the River Continuum Concept was developed, showed a surprising amount of diatoms in their guts (Rosi-Marshall et al., 2016). Still, macroinvertebrates in the Salmon River headwaters consumed more allochthonous C. In fact, macroinvertebrates also consumed about 80% allochthonous C in this study. These two perspectives are important to consider: consumption versus assimilation. Both processes are essential aspects of C cycling; however, each indicate different C fates. Carbon assimilation compared with available resources can be used to quantify trophic transfer efficiency, while assimilation subtracted from consumption quantifies egestion that is a measure of C recycling (i.e., carbon available in a different form to a range of organisms). Therefore, autochthonous C trophic transfer was relatively high compared to resource availability. Yet, allochthonous C consumption was relatively high and critical for C recycling. Lamberti and Steinman (1997) reviewed studies from 30 streams, mostly in the Northern Hemisphere, that indicated periphyton was an energetically important food resource in a range of stream sizes, biomes, and geographic areas. Furthermore, Finlay (2001) compiled macroinvertebrate δ 13C values from 70

streams worldwide that indicated functional groups other than scrapers and shredders were assimilating significant amounts of epilithic algae, particularly in streams with catchments greater than 10 km2 . Our results, confirming the importance of primary producers to macroinvertebrate production reflect that (1) algae are assimilated more efficiently than leaf litter because of higher nutritive value (e.g., lower C:N:P content) (Berg and Hellenthal, 1992), (2) algae have higher turnover rates that increase availability to macroinvertebrate consumers (Lamberti et al., 1989), and (3) algae are available year round even in streams with nearly closed deciduous canopy (Hamilton et al., 2004).

When we compared the amount of algal C available (calculated as gross primary production; GPP) to the total amount of algae consumed, we found that macroinvertebrates were consuming 4–36% of the total GPP [or about 8– 72% net primary production (NPP)]. This autochthonous contribution of C to the macroinvertebrates is greater than we anticipated and greater than in tropical forested streams (Neres-Lima et al., 2017; Rosas et al., 2020). The contribution of autochthonous C was likely relatively high in this stream and perhaps others because: (1) filtering taxa were likely ingesting recycled autochthonous C in the form of epilithic scrapers and sloughed particles (e.g., Strayer, 1988; Benke and Wallace, 1997), (2) scrapers were likely increasing algal turnover rates through efficient grazing, while keeping biomass low (e.g., Lamberti et al., 1995), and (3), estimation errors associated with both whole-reach estimates of GPP and NPP (e.g., Hall and Tank, 2003) and using published estimates of C assimilation by macroinvertebrates using measurements of production could under estimate C consumption (e.g., Morin et al., 1987). Our results highlight assumptions that must be made and the difficulty in estimating the contribution of autochthonous and allochthonous resources in systems with consumers ingesting a complex mixture of both types simultaneously.

When we compare the amount of leaf litter consumed by macroinvertebrate community compared to the amount available, consumption values ranged from 13 to 40% of available. Organic matter evidently was not limiting macroinvertebrate population growth on an annual scale, although it might have been limited in some seasons, as has been shown in other studies (e.g., Richardson, 1991; Dobson and Hildrew, 1992; Marks, 2019). The addition of large wood increased the percentage of GPP and coarse benthic organic matter consumed by 25 and 10%, respectively, for the macroinvertebrate community in the treatment reach relative to the control reach. Other studies have found that changes in microhabitat, such as increased leaf litter accumulations, may change the invertebrate community structure and increase production (e.g., Wallace et al., 1997a; Kobayashi and Kagaya, 2004). In our study, increased patches of inorganic coarse substrate and deposited leaf litter and fine sediment associated with the added wood seemed to have increased production by the shorter-lived taxa, such as Baetis, resulting in increased community-level secondary production (Entrekin et al., 2009). Longer-term increases

in organic matter retention are predicted to increase the relative importance of allochthonous organic matter to the food web and sustain greater production and trophic transfer in the future.

#### DATA AVAILABILITY STATEMENT

All macroinvertebrate datasets generated for this study are included in the article/**Supplementary Material**.

#### AUTHOR CONTRIBUTIONS

SE co-designed the experiments, analyzed the data, and was the lead manuscript writer. ER, JT, GL, and TH co-designed the experiments and provided critical input and edits to the manuscript.

#### FUNDING

The authors declare that this study received funding from provided by a grant from the USDA-NRI Managed Ecosystems Program (2003-35101-12871) to JT and GL. Additional funding to SE was provided by a Sigma Xi grant, a Bayer Fellowship, and a

#### REFERENCES

Warner-Lambert Fellowship from the University of Notre Dame. The funders were not involved in the study design, collection, analysis, interpretation of data, the writing of this article or the decision to submit it for publication.

#### ACKNOWLEDGMENTS

We thank Aileen Kelly and Brian Holland for assistance with sample processing and Dennis Birdsell and Jon Loftus for analytical support at the Center for Environmental Science at University of Notre Dame. We also thank personnel of the Ottawa National Forest, United States Department of Agriculture Forest Service for support throughout the study. Comments by David Janetski, Scott Tiegs, Barbara Downes, Richard Johnson, CA, and two reviewers improved the manuscript substantially.

#### SUPPLEMENTARY MATERIAL

The Supplementary Material for this article can be found online at: https://www.frontiersin.org/articles/10.3389/fevo.2020.00114/ full#supplementary-material

the results of an isotope addition experiment. J. North Am. Benthol. Soc. 23, 429–448.

Marcarelli, A. M., Baxter, C. V., Mineau, M. M., and Hall, R. O. Jr. (2011). Quantity and quality: unifying food web and ecosystem perspectives on the role of resource subsidies in freshwaters. Ecology 92, 1215–1225. doi: 10.1890/10-2240.1

**Conflict of Interest:** The funders were not involved in the study design, collection, analysis, interpretation of data, the writing of this article or the decision to submit it for publication.

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Copyright © 2020 Entrekin, Rosi, Tank, Hoellein and Lamberti. This is an openaccess article distributed under the terms of the Creative Commons Attribution License (CC BY). The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

# Drought-Induced, Punctuated Loss of Freshwater Mussels Alters Ecosystem Function Across Temporal Scales

#### Traci P. DuBose1 \*, Carla L. Atkinson2 , Caryn C. Vaughn1 and Stephen W. Golladay 3

*1 Oklahoma Biological Survey & Department of Biology, University of Oklahoma, Norman, OK, United States, 2 Department of Biological Sciences, University of Alabama, Tuscaloosa, AL, United States, 3 Jones Center at Ichauway, Newton, GA, United States*

#### Edited by:

*M. Eric Benbow, Michigan State University, United States*

#### Reviewed by:

*William Giles McDowell, Merrimack College, United States Brandon Barton, Mississippi State University, United States*

> \*Correspondence: *Traci P. DuBose<EMAIL_ADDRESS>
#### Specialty section:

*This article was submitted to Population and Evolutionary Dynamics, a section of the journal Frontiers in Ecology and Evolution*

Received: *22 December 2018* Accepted: *02 July 2019* Published: *18 July 2019*

#### Citation:

*DuBose TP, Atkinson CL, Vaughn CC and Golladay SW (2019) Drought-Induced, Punctuated Loss of Freshwater Mussels Alters Ecosystem Function Across Temporal Scales. Front. Ecol. Evol. 7:274. doi: 10.3389/fevo.2019.00274* Punctuated, mass mortality events are increasing for many animal taxa and are often related to climatic extremes such as drought. Freshwater mussels are experiencing increased mass mortality events linked to hydrologic drought. Because mussels play important functional roles in rivers it is important to understand the ecosystem effects of these die-offs. Here, we address how mass mortality events of mussels caused by drought may impact stream ecosystem function. We first present a conceptual model, based on the literature, of how mussel mass mortality should affect different ecosystem functions across various ecological time scales, from hours to decades. Next, we highlight two case studies of drought-linked, mussel-mass mortality events from rivers in the southern U.S. We then present the results of an experiment we performed quantifying the ecosystem effects of a punctuated mussel die-off. Finally, we combine our experimental results with field data from a recent mussel die-off to predict how mussel losses will influence ecosystem function. Based on the presented case studies, our mesocosm experiment, and our extrapolated nutrient pulse due to a mussel die-off, we conclude that stream ecosystems are extensively altered following mussel mass mortality events. Mussel loss is governed by drought severity, location within the river network, and species-specific drought tolerances. In the short term, decomposing carrion from mussel die-offs releases a large pulse of nutrients into the water which stimulates food web productivity. In the long term, the overall loss of mussel biomass, and the loss of functional traits as more sensitive species decline, leads to decreases in ecosystem function which may take decades to recover. Drought and human demand for water will make mussel die-offs more likely in the future and it is unlikely that drought sensitive species will recover without changes in water management and restoration of populations through mussel propagation. Our research provides an example of how the loss of an abundant, long-lived organism has cascading, and long-term impacts on ecosystems.

 These pulses are widespread and often result from climatic and environmental factors. Resource pulses can result from the mass die-offs of animals, such as 17-year cicadas, spawning salmon, and even wildebeest, and are increasingly recognized as important components of ecosystem function (Yang et al., 2008; Subalusky et al., 2017). Mass mortality events, or die-offs, are increasing in frequency across most taxa (Fey et al., 2015), thus it is important that we understand how these events affect ecosystem function (Baruzzi et al., 2018). In freshwater systems, unionid mussels play important structural and functional roles and are also experiencing increasing mass mortality events globally (Lydeard et al., 2004; Wenger et al., 2018) that are often linked to climatic events such as drought (Vaughn et al., 2015). Thus, they are a useful system for investigating the ecosystem effects of mass mortality events, particularly as related to environmental change (Fey et al., 2015).

Here, we address how mass mortality events of unionid mussels caused by drought may impact stream ecosystem function short-term and long-term. We first present a conceptual model, based on the literature, of how mussel mass mortality should affect different ecosystem functions across various ecological time scales, from hours to decades. Next, we highlight two case studies of drought-linked, mussel-mass mortality events from rivers in the southern U.S. We then present the results of an experiment we performed quantifying the ecosystem effects of a punctuated mussel die-off. Finally, we combine our experimental results with field data from a recent mussel die-off to predict how mussel losses will influence ecosystem function.

#### HOW DO MUSSEL MASS MORTALITY EVENTS IMPACT ECOSYSTEM FUNCTION? A CONCEPTUAL MODEL

Freshwater mussels (order Unionida, hereafter mussels or unionids) are sedentary mollusks that live burrowed in stream sediments where they filter the water and transfer energy and nutrients from the water column to benthos. Nutrients excreted and biodeposited by mussels stimulate instream microbial, primary, and secondary production and are even exported to riparian areas (Allen et al., 2012; Vaughn, 2018). Mussels provide habitat for other organisms through the biogenic structure of their shells and by changing hydrodynamic conditions at the sediment-water interface (Sansom et al., 2018b). They are longlived (6–100 years) with high native biodiversity in eastern North America (Williams et al., 1993), often live in high density, multispecies aggregations (hereafter mussel beds) that can persist in rivers for many decades and can make up most of the invertebrate biomass in many perennial rivers (Sansom et al., 2018a). Mass mortality of mussels has been linked to increasing drought, either from emersion (Atkinson et al., 2014) or from low dissolved oxygen and high temperatures associated with decreased water volume (Gagnon et al., 2004), as drying in streams is often accompanied by increased water temperatures and diel oxygen shifts (Mosley, 2015). How mussels respond to drought conditions depends on individual species' physiological tolerances, drought severity, and abiotic conditions (Gagnon et al., 2004; Golladay et al., 2004; Haag and Warren, 2008; Gough et al., 2012). When mass mortality of mussels occurs, their loss should influence stream ecosystem functions in a variety of ways across ecological time scales (**Figure 1**). We highlight these predicted effects below.

### Mussel Mortality Feedback Loop

Mussel species vary in their physiological tolerance and response to stress (Spooner and Vaughn, 2008). Species that are sensitive to low oxygen (hypoxia sensitive) or higher temperatures (thermally sensitive) are less likely to survive during a hydrologic drought, thus thermally tolerant and/or hypoxia tolerant mussels become the dominant species within the assemblage (Gagnon et al., 2004; Atkinson et al., 2014). Mussel soft tissue can decay within seven days, as shown below. Decomposing soft tissue releases a pulse of nitrogen and phosphorus into the water column and interstitial spaces (**Figure 1A**; Cherry et al., 2005; Atkinson et al., 2014). Depending on stream discharge, this nutrient pulse moves downstream over a few hours/days. Pore water can retain high nutrient concentrations longer, potentially exposing burrowing unionids to lethal nutrient concentrations (Cooper et al., 2005; Gough et al., 2012). After the loss of mussels, shifts in algal production, and turbidity are driven by stream discharge. If the stream becomes intermittent (a series of drying pools), turbidity will likely decrease due to increased sedimentation while algal blooms will likely form in stagnant areas (Mosley, 2015). During intermittence, the combination of reduced biofiltration and the release of nutrients into the water from mussel soft tissue decay encourages large algal blooms (Gagnon et al., 2004), which leads to high respiration rates at night, further reducing dissolved oxygen concentrations and stressing the remaining mussels, leading to additional mortality (**Figure 1B**). This cycle can also exacerbate feedback among deaths within mussel beds; algal blooms cause mortality in remaining hypoxia-intolerant bivalves, worsening algal blooms, and depressing dissolved oxygen, further stressing and eventually killing hypoxia-tolerant mussels. If the stream remains perennial, turbidity will increase through the addition of suspended solids from upstream and algal blooms become less likely. Algal blooms and/or increased turbidity can persist because of a reduction in biofiltration by freshwater mussels (**Figure 1C**).

#### Reduced Biofiltration

Immediately after a mass mortality event, biofiltration is greatly reduced in part to both residual stress on the remaining living mussels and biomass loss from mussel mortality (Vaughn et al., 2015). Biofiltration by the remaining mussels will gradually increase within the following days, but is likely to remain low until mussel biomass is replaced, unlikely for at least a decade (**Figure 1E**). Reduced biofiltration drastically increases the time required for the remaining mussels to filter a given

downstream values for 1991 in Vaughn et al. (2015) for the biofiltration and community structure data and on Atkinson and Vaughn (2015) for downstream nutrient excretion and storage information. \*Loss of nutrient capacitance dependent on mussel recruitment. Time increases to the left and the impact on mussels of four main ecosystem functions is described and depicted. Capital letters refer to distinct processes in the model, while lower case letters refer to sources: (a) Vaughn et al. (2015), (b) Atkinson and Vaughn (2015), (c) McDowell et al. (2017), (d) Gagnon et al. (2004), (e) Cherry et al. (2005), (f) Cherry et al. (2005), (g) Atkinson et al. (2014), (h) Atkinson et al. (2018), (i) Ilarri et al. (2018).

amount of water, reducing material exchange between the water column and benthos (Baustian et al., 2014; Vaughn et al., 2015).

### Reduced Nutrient Capacitance and Storage

Short-term nutrient storage in mussel soft tissue is greatly reduced through decomposition (Atkinson et al., 2014). As mussels filter feed, they act as "nutrient capacitors," accumulating, storing, and releasing energy and nutrients (carbon, nitrogen, and phosphorus) at different rates based on their age and species' traits (Strayer, 2014; Atkinson et al., 2018). Following massmortality, remaining mussel assemblages have lower abundance, age diversity, and species diversity, reducing their ability to filter seston and excrete nutrients. This reduced nutrient capacitance may result in longer nutrient spirals and more downstream transport of nutrients, likely due to an increase in nutrient uptake length (**Figure 1D**; Atkinson et al., 2014). Shells of deceased mussels lose ∼50% mass by 15 years through mechanical and chemical dissolution, which reduces the nutrient storage and shell habitat within the mussel bed as particle size becomes more homogeneous (**Figure 1E**; Atkinson et al., 2018).

### Changes in Habitat Provided by Shells

Live mussels and their spent shells physically modify the environment in streams, providing unique habitat for other organisms. Tissue decay potentially creates interstitial spaces within the substrate, which can be used by both macroinvertebrates and fish. Shells vary in shape and size across species and age and can accumulate in the sediment at different rates; thus, shell habitat can harbor variable macroinvertebrate communities depending on the shells species of origin (Bódis et al., 2014). While shells represent hard surfaces for macroinvertebrates, they dissolve over time. Shell dissolution is fastest in flowing waters with low calcium carbonate concentrations and thin, small shells dissolve faster than thicker, larger shells (Strayer and Malcom, 2007; Ilarri et al., 2015, 2019). While bivalve soft tissue decomposes quickly, shells persist for many decades (5–30 years), providing habitat for other stream biota (Strayer and Malcom, 2007; Ilarri et al., 2015; Atkinson et al., 2018). Over time, the benthos will be more homogenous as old shells dissolve and new relic shells are produced by a less diverse mussel assemblage ultimately altering benthic microhabitat characteristics and macroinvertebrate community structure (**Figure 1E**; Ilarri et al., 2018).

### Shifts in Community Composition and Ecosystem Function

During droughts, species sensitive to low oxygen (hypoxia sensitive) or higher temperatures (thermally sensitive) face greater risk of mass mortality leading to differential survival resulting in tolerant species becoming dominant within an assemblage, changing community structure (Gagnon et al., 2004; Atkinson et al., 2014). Surviving mussels may contribute to population recovery if conditions are suitable for reproduction. Most mussels have an ectoparasitic larval phase that requires a host fish (Barnhart et al., 2008). Drought concentrates fish into drying pools as they to attempt to escape harsh conditions or die due to increased biotic and abiotic stressors (Matthews and Marsh-Matthews, 2003; Lennox et al., 2019). While mussel reproduction is unlikely limited by host density (Haag and Stoeckel, 2015), different mussel species exhibit different host specificity and infection phenology (Barnhart et al., 2008). Thus, predicting the recruitment success is difficult due to unionid's unique life histories. As it takes mussels anywhere from 9 months to 10 years to reach sexual maturity, with most mussels reaching maturity around 4 years old (Haag, 2012), in die-off affected areas, biomass is unlikely to rebound for at least a decade. Future mussel assemblage structure is dependent on the surviving mussel assemblage, the surviving fish assemblage, and the recurrence frequency of droughts. If drought frequency decreases or remains constant, the mussel community could return to its former, pre-drought structure if no mussel species were extirpated from the river basin. If droughts increase in frequency and severity as projected in many regions (Palmer et al., 2008; USGCRP., 2017), we anticipate the mussel community will become dominated by tolerant mussel species (**Figure 1E**).

Mussel species with different temperature tolerances have different, temperature-dependent biofiltration and nutrient excretion rates. Thus, when the proportion of thermally sensitive vs. tolerant species in a mussel assemblage changes, this can impact ecosystem function (Spooner and Vaughn, 2008; Vaughn et al., 2015; Atkinson et al., 2018). For example, in rivers in southern Oklahoma, Actinonaias ligamentina is a thermally sensitive species with higher filtration and nutrient excretion rates at summer temperatures than other mussels in the assemblage. Because of its temperature intolerance, it also has a higher mortality rate during drought than other species. Thus, when this species is lost, the overall biofiltration (**Figure 1C**) and nutrient recycling capacity (**Figure 1F**) of the community is decreased for an extended time period, even if the biomass of other species remains stable (Vaughn et al., 2015). Further, mussel species excrete at different N:P ratios, and losses can also lead to shifts in excretion stoichiometry (**Figure 1F**; Atkinson et al., 2018). These changes can cascade through the food web impacting algal, macroinvertebrate (Novais et al., 2017), and even fish dynamics.

### TWO CASE STUDIES OF DROUGHT-DRIVEN MUSSEL LOSSES IN THE SOUTHERN UNITED STATES

Drought-induced mussel mass mortality events have been documented for two diverse, well-studied river systems in the southern U.S., the Lower Flint River in Georgia and the Kiamichi-Little River system in Oklahoma (**Figure 2**). These case-studies allow for a deeper understanding of how drought affects mussel assemblages, the subsequent changes in stream ecosystems, and the potential recovery time for mussel assemblages.

#### Lower Flint River, Georgia

The lower Flint River of southwestern Georgia experienced an extended period of below normal rainfall from 1999 until 2013. This period included three multiyear droughts that were classified as severe/exceptional. The summer of 2000 was particularly disastrous for unionids as streams in the region experienced unprecedented low flows and transitioned from perennial to intermittent. Forty-six historically-species-rich sites in lower Flint tributaries were surveyed in 1999 prior to drought onset, and a subset of these were resampled in 2000 (Gagnon et al., 2004). Stream drying had not been previously observed but became common during subsequent growing seasons (Rugel et al., 2012). Stream flow was essential for maintaining dissolved oxygen concentrations within the tolerances of freshwater mussels; dissolved oxygen between 5 and 3 mg L−1 resulted in 24% mortality and when dissolved oxygen fell below 3 mg L−1 up to 76% of mussels died (Gagnon et al., 2004; Golladay et al., 2004). These degraded physicochemical conditions differentially impacted species within the mussel communities. Riffle specialists in medium-sized streams suffered the highest mortality, while drought-tolerant, small stream species and species in larger tributaries whose habitat was buffered from drought conditions fared better (Gagnon et al., 2004). In medium-sized streams, community composition shifted toward more generalist species from riffle specialists (Gagnon et al., 2004).

The lower Flint River has been surveyed since this drought. Through 2013, there was little evidence of recovery from mortality associated with the 1999–2001 drought (Smith et al., 2015). During most summers since the initial drought, conditions were stressful, likely preventing the reproduction of surviving unionids. Reproduction of mussel populations, as evident from observation of juvenile mussels, was not apparent until rainfall approached average levels (2013– 2015) (Smith et al., 2015). The extended period of below normal rainfall (1999–2013) and subsequent below normal stream flows were likely exacerbated by anthropogenic water withdrawal; the mid reaches of tributaries of the lower Flint cross the Dougherty Plain physiographic district, which is a recharge area for a heavily developed agricultural water source (Golladay et al., 2004). This case study provides evidence that mussel biomass might not recover from mass mortality events for over a decade and that anthropogenic and climate alterations can alter stream benthic communities for extended periods.

#### Kiamichi and Little Rivers, Oklahoma

The Kiamichi and Little Rivers in southeastern Oklahoma are adjacent, major tributaries to the Red River. This region experienced a period of exceptional drought during 2011–12 where the Kiamichi River experienced 84 days of no flow (defined

as discharge < 0.01 m3 s −1 ) and 36 weeks of extreme low flow, defined as flows below the 10th percentile of flow frequency (Atkinson et al., 2014). The Little River, and its major tributary the Mountain Fork River, experienced 39 and 40 weeks of extreme low flow, respectively (Atkinson et al., 2014).

study section are named. Code from the R package *hydroMap* (DeCicoo and Blodgett, 2017).

These severe drought conditions led to a mass mortality event as mussels became isolated in shallow drying pools or emersed. Mussel losses and their effects on ecosystem function were documented in two related studies. Atkinson et al. (2014) sampled mussels at three sites before a drought (2010) and at the end of the drought (2012) and assessed changes in mussel abundance and mussel-provided nutrient cycling and storage. Sixty percent of unionids died during the drought, but thermally sensitive species had a higher mortality rate, resulting in a community shift toward more thermally tolerant species. In the second study, Vaughn et al. (2015) compared mussel biomass and ecosystem services (biofiltration, nutrient cycling, and nutrient storage) at four sites in the Kiamichi River across several decades (1991, 2004, 2011). 1991 was a wet period and 2004 and 2011 were drought periods with significant mussel losses. They found that mussel biomass decreased over 60% across these sites and that ecosystem function losses mirrored the biomass losses. The sites experienced mussel biomass losses of ∼28% and corresponding declines in nitrogen recycling (22%), phosphorus recycling (15%), and ∼30% declines in areal storage of nitrogen and phosphorus (Vaughn et al., 2015).

### AN EXPERIMENT QUANTIFYING THE EFFECTS OF A MUSSEL DIE-OFF ON ECOSYSTEM FUNCTION

While case studies have documented how native mussel communities change after mass mortality events, no studies have experimentally demonstrated how these losses impact ecosystem function. With our conceptual model and case studies in mind, we designed an experiment to measure ecosystem function changes that occur following a punctuated mussel mass mortality event. We conducted a mesocosm experiment at the University of Oklahoma Biological Station in the summer of 2018 where we induced mussel mortality and measured effects on ecosystem structure (water column nutrient concentrations and algal abundance) and ecosystem function (decomposition rates and ecosystem metabolism) over time. We predicted that decaying mussel tissue would increase nutrient concentrations, which would stimulate both algal growth and microbial respiration and decomposition.

#### Mesocosm Experiment Methods

We used 18, 1.52 m diameter, 946 L circular tanks to simulate drying stream pools. Each mesocosm (tank) was lined with ∼15 cm of gravel (1:1 ratio of 10 and 38 mm diameter gravel). We had 9 control mesocosms with no mussels and 9 mesocosms containing 31 mussel individuals, to replicate a natural mussel community in the region. Each mussel treatment mesocosm contained 13 Actinonaias ligamentina, 9 Cyclonaias pustulosa, 5 Amblema plicata, two Tritogonia verrucosa, one Lampsilis cardium, and one Plectomerus dombeyanus. This represented a low, but natural density of mussels (11.9 mussels/m2 ) and reflected the freshwater mussel community of the upper Kiamichi River (Atkinson et al., 2012).

We describe sampling events and present our results relative to the day of the mussel mass mortality event: negative values indicate days before and positive values days after the mussels died. We filled mesocosms with water 12 days (day−45; see **Supplementary Table 1** for sampling dates and measurements) before adding mussels to allow the mesocosms to be naturally colonized by algae and macroinvertebrates. On day−33, we added mussels. On day−14, we introduced 10 largemouth bass (Micropterus salmoides; mean standard length = 95 mm, SD = 16 mm) to simulate how fish are concentrated in drying pools during early periods of drought. We removed the fish on day−2 to simulate their movement downstream during the drought as drying pools became too stressful for them (Magoulick and Kobza, 2003). Fish may have impacted water column nutrients on day−3 but had little impact on mesocosm nutrient concentrations and algal abundance (see Results below). We induced a punctuated mass mortality event on 2 July 2018 (day 0) by sacrificing the mussels in 5 of the 9 mussel mesocosms, while maintaining the 9 non-mussel controls. To produce the carrion for this stage of this experiment, we cut the adductor muscles of 155 mussels. We returned the mussel carrion to the mesocosms to allow for natural decomposition of soft tissue.

We sampled mesocosms 3 times before and 4 times after the mass mortality event, resulting in the following sampling days:−20,−15,−3, 4, 11, 25, and 39. On each sampling day we measured dissolved oxygen (DO), conductivity, and water temperature at midday (**Supplementary Table 2**). To measure water column phosphorus and nitrogen concentrations, we collected 20 mL filtered water samples (Grade F, 0.7µm pore size, Sterlitech Kent, Washington) and froze them for subsequent nutrient analysis. We lost nutrient samples from day 11, thus we resampled the mesocosms on day 18. We filtered two water samples (Grade A, 1.6µm pore size, Sterlitech Kent, Washington) from each mesocosm to quantify water column chlorophyll a. Filters were frozen for later chlorophyll estimation. On day−33, we placed six 7.6 cm2 clay tiles with a 27.5 mm2 fritted glass disc attached with silicone (LECO cover crucible AL P 1000; GE Silicone 1∗ All Purpose) on the substrate surface to allow algal colonization for sampling of benthic algal production. We removed two glass fritted discs on each sampling day and froze them for later estimation of benthic algal biomass.

We quantified soluble reactive phosphorus (SRP) with the colorimetric method (Murphy and Riley, 1962; Stainton et al., 1974; EPA Method 365.3) and ammonium (NH4-N) using the phenol method (5.2.6 EPA Method 350.1; ASTM., 2012) for the filtered water samples. To measure chlorophyll a concentration, we cold-extracted water column (filters) and benthic (fritted discs) samples with acetone and measured the extractant spectrophotometrically with a correction for pheophyton (ASTM., 2012).

We followed Tank et al. (2017) to measure ecosystem metabolism as gross oxygen production on days 4, 11, 25, and 39. We quantified ecosystem metabolism by measuring dissolved oxygen production and respiration in light and dark cycles, respectively, on the glass fritted discs in 50 mL centrifuge tubes. We measured dissolved oxygen (HACH HQ40d multiple parameter meter, Loveland, Colorado) to estimate initial oxygen concentrations. We placed fritted disks in centrifuge tubes filled with the respective mesocosm's water and sealed the tubes. After allowing the discs to metabolize for an average of 1.75 h (SD = 0.23 h) in a common mesocosm, we re-measured the dissolved oxygen within the tubes. We then removed the water and repeated the filling process with the same tube/glass fritted disc pair. After filling the tubes, we immediately placed them in the dark in a common mesocosm for an average of 2.62 h (SD = 0.47 h). We then measured final dissolved oxygen and collected and froze the discs. Gross primary production (mg DO cm−1 h −1 ) was calculated from the addition of net primary productivity (difference in DO in the light cycle) and the absolute value of the respiration (DO difference in the dark cycle).