Linking Aquatic Biodiversity Loss to Animal Product Consumption: a Review

MSc Biological Sciences Freshwater and Marine Biology

Literature Review

Linking aquatic biodiversity loss to animal product consumption: A review

Michelle Pena-Ortiz 11567791

July 2021 12 ETCS

Supervisor & Assessor: Dr. Brian Machovina 1 Examiner: Dr. Harm van der Geest 2

1 College of Arts, Sciences, and Education – Florida International University 2 Institute for Biodiversity and Ecosystem Dynamics - Universiteit van Amsterdam

Table of Contents Abstract ...... 3 1. Introduction ...... 5 2. Methodology ...... 8 3. Pressures ...... 8 3.1. Habitat alteration ...... 9 3.1.1. Riparian zones ...... 10 3.1.2. Channelization ...... 12 3.1.3. Water extraction ...... 13 3.2. Climate change ...... 13

3.2.1. CH4 ...... 14

3.2.2. N2O ...... 16

3.2.3. CO2 ...... 17 3.3. Pollution ...... 19 3.3.1. Nutrients ...... 20 3.3.2. Sediment ...... 21 3.3.3. Antibiotics ...... 22 3.3.4. Estrogenic compounds ...... 23 3.4. Alien species ...... 23 3.5. Overexploitation...... 24 4. State ...... 26 4.1. Habitat alteration ...... 26 4.2. Climate change ...... 28 4.2.1. Effects of increased in-water temperatures on aquatic biodiversity ...... 28 4.2.1. Effects of altered precipitation on aquatic biodiversity ...... 30 4.3. Pollution ...... 31 4.4. Alien species ...... 34 4.5. Overexploitation...... 36 4.6. Multi-stress ...... 37 5. Potential responses ...... 38 5.1. Production ...... 38 5.2. Consumption ...... 40 5. Conclusions ...... 43 Literature cited ...... 44

Abstract

Animal agriculture, which accounts for 83% of global farmland, but produces only 18% of global calories, is responsible for the majority of agriculture-related environmental degradation. The pressures that animal product consumption place on the environment include habitat alteration, climate change, pollution, invasive species introductions, and overharvesting of wild animals. These pressures are the five main drivers of biodiversity loss and for most ecosystems, each pressure is currently either constant or growing. It has recently been established that aquatic (freshwater) ecosystems are particularly vulnerable to these pressures yet remain less studied and less protected than terrestrial and marine ecosystems. Although it is well known that animal product consumption is a leading driver of the aforementioned pressures, and aquatic biodiversity is declining as a result of these pressures, the link between animal product consumption and aquatic biodiversity loss has not been thoroughly explored. Therefore, the aim of this literature review was to evaluate how animal product consumption contributes to aquatic biodiversity loss by analyzing the contribution of animal agriculture to the five main drivers of biodiversity loss and investigating how aquatic biodiversity is affected by each of these drivers. Moreover, this review proposed potential responses that would help mitigate the effect of these pressures on aquatic biodiversity. This review found that there is a strong link between animal product consumption and aquatic biodiversity loss, and that animal agriculture is likely a primary driver of aquatic biodiversity loss. Research on potential responses demonstrated that decreasing or eliminating consumption of animal products was the most impactful strategy for alleviating pressures placed on aquatic biodiversity when compared to supply-side interventions, such as increasing efficiency. More specifically, global adoption of plant-based diets could reduce agricultural land-use by 76%, which is equivalent to an area more than three times the size of the

United States, including Alaska, Hawaii, and Puerto Rico, as well as result in 358-743 GtCO2 removal via carbon sequestrations through restoration, which is roughly equivalent to the past 9 to 16 years of fossil fuel emissions.

Keywords: animal products, agriculture, aquatic biodiversity loss, meat consumption, plant- based diets

Figure 1. Graphical abstract depicting the five pressures of animal product consumption on aquatic biodiversity in a river (left), and aquatic biodiversity in the absence of these pressures, for reference (right). Depictions of pressures are numbered as follows: (1) Habitat alteration i.e., degraded riparian buffer, (2) Climate change due to release of GHGs, (3) Pollution from nutrients, sediment, and pharmaceuticals, (4) Alien species escaping from aquaculture, and (5) Overconsumption of wild freshwater animals.

1. Introduction

It is widely known that global agriculture is a predominant driver of various forms of environmental degradation (Campbell et al., 2017; Foley et al., 2005; Steinfeld et al., 2006). Animal agriculture, which accounts for 83% of global farmland, but produces only 18% of global calories (Poore & Nemecek, 2018), is responsible for the majority of agriculture-related environmental degradation (Clark & Tilman, 2017; Machovina et al., 2015; Springmann et al., 2018; Willet et al., 2019). In addition to animal agriculture, the direct consumption of animals via harvesting (e.g., hunting and fishing) also contributes to environmental degradation by altering trophic interactions and decreasing biodiversity (Machovina et al., 2015). Despite the urgent need to curtail the pressures that animal consumption places on the environment, it is estimated that the demand for meat and dairy will increase by 68% and 57%, respectively, by 2030 (McLeod, 2011).

Myriad studies have demonstrated that animal products with the lowest impact on the environment, such as poultry and eggs, still exceed average environmental impacts of crops across indicators such as greenhouse gas (GHG) emissions, land use change, nutrient pollution, and energy use (Clark & Tilman, 2017; Poore & Nemecek, 2018; Ripple et al., 2014; Swain et al., 2018; Willett et al., 2019). Animal products most commonly include beef, pork, chicken, fish, dairy, and eggs, which are produced in conventional systems such as Confined Animal Feeding Operations (CAFO), pasture or free-range systems, aquaculture ponds, and/or wild harvest. The key reason for the significantly larger environmental impacts of animal products, compared to crop products, is the inherent inefficiency of energy transfer between trophic levels when converting plant calories into animal calories (Cassidy et al., 2013; Swain et al., 2018). Therefore, when addressing animal product consumption’s effects on the environment, it is important to consider not only the resources utilized directly by the animals, but the feed crops as well. Studies show that the ratio of animal product calories to feed calories is on average 7-10%, yet conversion efficiency is dependent on the animal product (Godfray et al., 2010; Sherpon et al., 2016). Sherpon et al. (2016) found that conversion efficiency between crop and animal products in the

US can ranges from 3% to 17%, suggesting that using edible crops to feed animals is an inefficient way to provide food for humans (Figure 2).

Figure 2. A Sankey flow diagram of the US feed-to-food caloric flux from the three feed classes (left) into edible animal products (right). On the right, parenthetical percentages are the food - out/feed-in caloric conversion efficiencies of individual livestock categories. From Sherpon et al. (2016)

The pressures that animal product consumption place on the environment include habitat alteration, climate change, pollution, invasive species introductions, and overharvesting of wild animals (Allan et al., 2005; Clark & Tilman, 2017; De Silva et al., 2009; Poore & Nemecek, 2018; Ripple et al., 2014; Swain et al., 2018; Willett et al., 2019). According to the Millennium Ecosystem Assessment (2005), these pressures are the 5 main drivers of biodiversity loss, and for most ecosystems, each pressure is currently either constant or growing. This suggests that animal product consumption is a major driver of biodiversity loss, and potentially the leading cause of modern species extinctions (Machovina et al., 2015). It has recently been established that aquatic (freshwater) ecosystems are particularly vulnerable to these pressures yet remain less studied and less protected than terrestrial and marine ecosystems (MEA, 2005; WWF, 2020). More 6 specifically, the pressures of habitat alteration, climate change, pollution, and alien species introductions have been found to be rapidly increasing and have a “very high” effect on biodiversity in freshwater ecosystems, while overharvesting remains stable, and has a “moderate” impact on freshwater biodiversity (MEA, 2005). Aquatic species populations are declining at a rate of 4% a year, resulting in a total loss of 83% since 1970, according to an analysis of 3,741 monitored populations in the Freshwater Living Planet Index (WWF, 2020). Other studies have found that almost one in three freshwater species are at risk of extinction, with all taxonomic groups showing a higher risk of extinction than terrestrial ecosystems (Collen et al., 2014). Although it is well known that animal product consumption is a leading driver of the aforementioned pressures, and aquatic biodiversity is declining as a result of these pressures, the link between animal product consumption and aquatic biodiversity loss has not been thoroughly explored.

Therefore, the aim of this literature review was to evaluate how animal product consumption contributes to aquatic biodiversity loss by analyzing the contribution of animal agriculture to the five main drivers of biodiversity loss and investigating how aquatic biodiversity is affected by each of these drivers. Analyzing the effects of animal product consumption on aquatic biodiversity loss should give new insight into potential responses societies can adopt to mitigate this pressing environmental issue. The following questions have been raised to address the aim of this review: 1. What pressures are being placed on aquatic biodiversity as a result of animal product consumption? 2. What is the state of aquatic biodiversity as a result of these pressures? 3. What are potential responses to mitigate aquatic biodiversity loss driven by animal product consumption?

To answer these questions, this literature review will remain specific to the United States (US), a megadiverse country, where meat consumption is three times greater than the global average (Poore & Nemecek, 2018). The structure of this review is loosely modeled after the FAO’s pressure-state-response (PSR) framework (OECD, 1999).

2. Methodology

This literature review focused on acquiring available data that demonstrated the pressures that animal agriculture places on aquatic biodiversity, data that showed how aquatic biodiversity is affected by these pressures, and lastly, data that illustrated potential solutions to combat these pressures and improve the current state of aquatic biodiversity as a result of these pressures. Literature was primarily collected using academic search engines GoogleScholar, ScienceDirect, and University of Amsterdam’s publication database CataloguePlus. Relevant literature was found using key terms “animal agriculture”, “livestock grazing”, “USA CAFOs”, “livestock nutrient runoff”, “plant-based diets”, “overfishing freshwater US”, “antibiotics CAFO”, “estrogenic pollution livestock”, “aquatic invertebrate livestock”, and “climate change”. National agriculture statistics were retrieved by accessing governmental databases of the Environmental Protection Agency (EPA) and United States Department of Agriculture (USDA). This review aimed to synthesize existing scientific literature and draw conclusions based on data published by other authors, rather than contribute new findings. This review also offered conclusions based on available data and provided insight into knowledge gaps that could be addressed in future studies.

3. Pressures

The pressures that animal product consumption place on the aquatic biodiversity include habitat alteration, climate change, pollution, alien species introductions, and overexploitation of wild animals. The following sections will focus on the contributions of animal agriculture to each of these pressures. It is important to note that numerous studies demonstrate that animal products differ in their contribution to these pressures.

Figure 3. Overview of environmental pressures of animal agriculture. (A–D) Environmental performance of the key livestock categories in the US diet, jointly accounting for >96% of animal - based calories. Authors report performance in resources required for producing a consumed Mcal (1 Mcal = 1000 kcal, roughly half a person’s mean daily caloric needs). For comparison, resource demands of staple plants potatoes (denoted p), rice (r), and wheat (w) are denoted by arrows above A–D. E displays actual US consumption of animal-based calories. Values to the right of the bars denote categories’ percentages in the mean US diet. The demands of beef are larger than the figure scale and are thus written explicitly next to the red bars representing beef. Error (uncertainty) bars indicate SD. In A, for beef and dairy, demand for pastureland is marked with white hatching, and a vertical line separates demand for cropland (to the left), and processed roughage land (to the right). From Eshel et al. (2014).

3.1. Habitat alteration Animal product production (including feed crop production) takes up approximately 43% of all ice and desert-free land on earth, making it the single largest anthropogenic land use on the planet (Poore & Nemecek, 2018). In the US, grazing land for cattle accounted for approximately 35% - or 798 million acres - of total land area in 2012, making it the single largest land use type in the country (Bigelow & Borchers, 2017). Moreover, cropland accounts for approximately 17% - or 392 million acres – of total land in the US (Bigelow & Borchers, 2017), 40% of which is dedicated to animal feed (Eshel et al., 2014). Pasture-raised cattle typically account for the most terrestrial land use in the US (Figure 3), whereas in pig and chicken systems, nearly all associated land use is for feed crop production (de Vries & de Boer, 2010). Terrestrial pasturelands and feed croplands alter adjacent aquatic habitats by degrading riparian zones that are important for regulating shade, temperature, and organic matter inputs (Belsky et al., 1999; Goss & Roper, 2018; Jones et al., 2001; Wetzel, 2001). Aside from terrestrial land use change, animal product

9 production directly alters aquatic ecosystems via drainage alteration and/or channelization of streams and rivers that is carried out to facilitate feed crop yields (Lau et al., 2006; Raborn & Schramm, 2003; Sullivan et al., 2004). Water extraction for feed crops and pasture lands modifies aquatic ecosystems by reducing flow (Richter et al., 2020).

3.1.1. Riparian zones Riparian zones are considered the interface between terrestrial and aquatic ecosystems, and provide aquatic ecosystems with shade, allochthonous material (food), and habitat, while at the same time maintaining bank stability. Livestock grazing has been shown to be a key factor in the continued degradation of riparian zones (Batchelor et al., 2014; Behnke, 1978; Platts, 1982), particularly in the western US (Belsky et al., 1999; USDI, 1994). According to Armour et al., (1994), grazing by livestock has damaged 50% of riparian zones in arid regions of the western US, whereas US Department of Interior (1994) found that livestock grazing had damaged 80% of streams and riparian zones in the arid west. Myriad studies demonstrate that cattle access to streams results in loss of bank vegetation and trampling of streambanks (Batchelor et al., 2014; Bryant et al. 1972; Goss & Roper, 2018; Kauffman et al., 2004; Trimble & Mendel, 1995), which degrades stream water quality and in-stream habitats (Armour et al., 1994; Kauffman & Krueger, 1984).

Grazing cattle decrease riparian vegetative biomass and therefore overhead cover, as well as alter vegetative composition and diversity, which has profound impacts on aquatic habitats (Bryant et al. 1972, Chapman & Knudsen 1980; Evans & Krebs 1977; Kauffman et al., 2004; Knoph & Cannon 1982). In a study done on Rock Creek in Montana, Meehan & Platts (1978) observed that streamside riparian cover was 76.4% greater in ungrazed areas compared to grazed areas. Loss of riparian vegetative cover has been shown to increase average in-stream water temperatures, as well as decrease allochthonous inputs (Delong & Brusven, 1994; Sizer, 1992; Pozo et al., 1997; Jones et al., 2001; Wetzel, 2001). Previous studies have observed that lack of vegetative covering in cattle grazing areas can increase average in-stream water temperatures by 5-7°C (Nussle et al., 2015; Zoellick, 2004). Van Velson (1979) discovered that average water

10 temperatures decreased from 24°C to 22°C one year after cattle exclusion from a creek in Nebraska. Moreover, decreases in allochthonous material associated with lack of riparian cover decreases C/N ratios in lakes and streams, since the supply of C is degraded. Higher C/N ratios in aquatic habitats are associated with higher food web stability (Lu et al., 2014; Rooney & McCann, 2012), and as C/N ratios decrease, there is a shift in detrital processing and nutrient levels (Delong & Brusven, 1994; Sizer, 1992; Pozo et al., 1997). Associated changes in detrital processing in upstream reaches of a stream could affect downstream trophic dynamics via reduction of transported fine particulate organic matter (Cuffney et al., 1990).

Trampling of streambanks by cattle causes bank slumping (Figure 4), which eventually leads to stream channels becoming wider and shallower (Batchelor et al., 2014; Goss & Roper, 2018; Herbst et al., 2012; Knapp & Matthews, 1996; Trimble & Mendel, 1995). Reductions in-stream width/depth ratio create a larger stream surface area that is exposed to solar radiation, thus further exacerbating the increase in average stream water temperature that occurs from loss of vegetative overhead cover (Herbst et al., 2012; Kauffman et al., 1983). Additionally, trampling of streambanks, coupled with the loss of riparian vegetation, results in decreased infiltration potential of soil, leading to higher peak flows in streams (Behnke, 1978; Belsky et al., 1999). Typically, water that percolates into the ground moves through sub-soil and seeps into stream channels through the year, creating perennial flows. However, as cattle compact soil and degrade vegetation, less rainwater infiltrates the soil and instead flows overland into streams, creating peak flows (Belsky et al., 1999). Trimble & Mendel (1995) estimated that peak storm runoff from a 120-hectare basin in Arizona would be 2-3 times higher when riparian areas were heavily grazed compared to lightly grazed. High intensity flows in grazed areas further alter stream morphology by incising and/or eroding stream banks, and deepening channels (Belsky et al., 1999; USDI, 1994). Furthermore, the interruption of perennial flows in riparian areas results in aquatic ecosystems receiving less water during late-season flows (Kovalchik & Elmore, 1992; Li et al., 1994; Ponce & Lindquist, 1990). Degradation of riparian zones also contributes to increased sedimentation and nutrient loads in aquatic habitats, reviewed in later sections of this paper.

Figure 4. Erosion pins (EP) record a bank failure due to cattle access. The numbers are obsolete and refer to tables in Peppier & Fitzpatrick (2005). The yellow dashes mark the riparian block that failed. Image taken from Peppler & Fitzpatrick (2005).

3.1.2. Channelization Stream channelization is a common practice in US croplands, because it can reduce the risk of crop loss from excess water stress and allows farmers more control over harvest conditions (Blann et al., 2009; Spaling & Smit, 1995). Although most literature on this topic does not specify these channelized regions as “feed crop agriculture”, it is important to discuss channelization in relation to animal agriculture, as feed crops make up 40% of crop production in the US (Eshel et al., 2014). For example, the Midwest produces 33% of the world’s corn (FAO, 2017), with most of this corn being used as the main energy source in livestock feed (WAOB, 2020). Many of the streams in agricultural regions of the Midwest have been exposed to severe and repeated channelization (Edwards et al., 1984; Emerson, 1971). Channelization of streams negatively

12 modifies crucial habitat parameters such as pool/glide quality, riffle/run quality, channel morphology, and substrate (Lau et al., 2006; Raborn & Schramm, 2003; Sullivan et al., 2004). Unchanneled streams harbor a variety of velocities, depths, widths, and substrates as a result of natural stream meandering, producing well developed and interspersed riffle and pool habitats (Aadland, 1993; Gorman & Karr, 1978; Scarnecchia, 1988). Channelization causes homogenous stream width and depth, making it difficult to distinguish between riffle and pool habitats that are important for aquatic biodiversity (Gorman & Karr, 1978; Smiley & Gillespie 2010).

3.1.3. Water extraction Feed crop production requires approximately 45 billion m3 of water, totaling about 27% of the total national irrigation use (Eshel et al., 2014). Almost 80% of irrigated feed crops are fed to beef cattle, and approximately 15% are fed to dairy cattle, making cows the largest consumers of irrigated feed crops in the US (Eshel et al., 2014; Richter et al., 2020) (Figure 3). According to Richter et al., (2020), irrigation of cattle-feed crops is the single largest consumer at both regional and national scales in the US, with 17 western states producing nearly all the feed crops that require irrigation due to the arid nature of these ecosystems (Richter et al., 2020; Schaible & Aillery, 2012). More specifically, cattle-feed irrigation is the leading driver of flow depletion in one-third of all western US sub-watersheds, and more than one-quarter of western rivers in the US are depleted by over 75% during summer months, with cattle-feed irrigation being the largest water use in more than half of these rivers (Richter et al., 2020). River flow depletion due to cattle-feed irrigation occurs when surface water is extracted, or shallow groundwater near rivers is pumped. Water extraction has the potential to exacerbate other adverse anthropogenic impacts on aquatic habitats such as water pollution, decreased oxygen levels, and opportunity for alien species of fish to outcompete native species (Poff & Zimmerman, 2010; Richter et al., 1997).

3.2. Climate change Climate change is primarily caused by the release of greenhouse gases into the atmosphere (IPCC, 2013). Globally, livestock farming accounts for 14.5% (7.1 gigatons CO2-eq per annum) of total

13 anthropogenic GHG emissions (Gerber et al., 2013) and is the primary anthropogenic source of methane (CH4) and nitrous oxide (N2O) emissions, producing 44-50% and 53-60%, respectively (Gerber et al., 2013; Smith et al., 2007). The global livestock sector’s emissions are comprised of

44% CH4, 29% N2O, and 27% CO2 (IPCC, 2007), with dairy and beef cattle contributing approximately 65-77% of these emissions (Gerber et al., 2013; Herrero et al., 2013). Pigs, chickens, and small ruminants contribute much less to emission levels, with each representing between 7-10% of sector emissions (Gerber et al., 2013). In 2019, beef cattle accounted for 72% of total national livestock CH4 (EPA, 2021). In the US, fertilizer application and manure management make up a significant portion of national N2O emissions, whereas crop feed production, agricultural land use change, and crop feed transportation make up a significant portion of national CO2 emissions (EPA, 2021). GHG emissions has resulted in an increase of 1°C above pre-industrial levels (IPCC, 2013). This anthropogenic warming alters the water cycle, and therefore, spatial and seasonal precipitation patterns (IPCC, 2013). This change in air temperature has substantial impacts on aquatic ecosystems, including increasing in-stream water temperatures and altering in precipitation patterns, which lead to periods of floods and drought (Knouft & Ficklin, 2017).

3.2.1. CH4

Methane is estimated to have a global warming potential (GWP) of 28-36 CO2-eq over 100 years

(CO2 is used as a reference and therefore has a GWP of 1) and stays in the atmosphere for approximately 10 years (EPA, 2021; IPCC, 2013). It is important to note that GWP calculations differ. The EPA uses a 100-year GWP of 25 CO2-eq for the national GHG inventory, based on calculations from the 2007 IPCC report (EPA, 2021; IPCC, 2007). The higher GWP of CH4 compared to CO2 is attributed to the fact that CH4 absorbs considerably more energy and is a precursor to ozone, which is another GHG (EPA, 2021; IPCC, 2013). In the US, animal agriculture is the largest anthropogenic contributor of CH4, primarily via enteric fermentation and manure storage systems.

Enteric fermentation is responsible for most anthropogenic CH4 emissions in the US, representing

27.1% of total anthropogenic CH4 in 2019 (EPA, 2021). In ruminants (e.g., cows, goats), enteric fermentation is the process of microbes decomposing and fermenting plant materials such as starch, cellulose, and sugars in the digestive track or rumen. Enteric methane is a byproduct of this process and is expelled from the animal via burping (Gerber et al., 2013; Ripple et al., 2014). Nonruminant, or monogastric, animals such as pigs and chickens produce negligible enteric methane emissions in comparison (Ripple et al., 2014). The large contribution of enteric fermentation to national CH4 emission levels can be explained by the findings of Eshel et al., (2014), which show that dairy and beef are the two most consumed animal products in the US (kcal/person/day), accounting for 12% and 7% of all consumed calories (Figure 3). In 2019, enteric fermentation was responsible for emitting 178.6 million metric tons (MMT) CO2-eq, representing an increase of 13.9 MMT CO2-eq. since 1990 (EPA, 2021). This increase in emissions from 1990 to 2019 follows the trend of increases in beef and dairy cow populations (EPA, 2021).

th Manure management was the 4 largest contributor of anthropogenic CH4 in the US, representing 9.5% of total CH4 emissions in 2019 (EPA, 2021). Manure that is handled as a liquid or slurry is typically stored in an integral tank that is under a confinement facility or flushed into an external anaerobic waste lagoon or pond (EPA 2020). The anaerobic decomposition that takes place in these systems breaks down volatile solids in livestock manure, which produces CH4 (EPA,

2020; Gerber et al., 2013; Leytem et al., 2017). The EPA (2021) estimated that CH4 emissions from livestock manure management were 62.4 MMT CO2-eq. in 2019, representing an increase of 25.3

MMT CO2-eq. since 1990, primarily due to increased concentrations of dairy and swine production (EPA, 2021). In 2019, CH4 emissions from dairy cattle manure produced 32.0 MMT

CO2-eq, followed by swine manure, which produced 23.1 MMT CO2-eq. (EPA, 2021; Leytem et al.,

2017). Poultry and beef cattle manure produced 3.6 MMT CO2-eq and 3.4 MMT CO2-eq, respectively, in 2019 (EPA, 2021). This large discrepancy in CH4 manure management emissions is primarily due to the increased concentration of dairy cows and pigs in confinement facilities. Despite the national dairy cow count decreasing since 1990, the industry has become more

15 concentrated in certain states, resulting in higher density confinement facilities that utilize liquid- based management systems to flush or store manure (EPA, 2021).

3.2.2. N2O

Nitrous oxide is estimated to have a GWP of 265-298 CO2-eq over 100 years and remains in the atmosphere for more than 100 years, on average (IPCC, 2007; Steinfeld et al., 2006). The significantly higher GWP of N2O compared to CO2 is attributed to the fact that it is significantly more effective at absorbing thermal radiation and plays a role in degrading stratospheric ozone (Poffenbarger et al., 2018). Nitrous oxide is primarily produced as an intermediary between nitrification and denitrification, which are anthropogenically altered via agriculture (EPA, 2020;

EPA, 2021). The livestock sector in the US primarily contributes N2O through application of manure on crops and pastures and manure management and storage (EPA, 2020; EPA, 2021).

Agricultural soil management is responsible for most anthropogenic N2O emissions in the US, representing 75.4% of total anthropogenic N2O emissions, and 5.3% of total GHG emissions in 2019 (EPA, 2021). According to the EPA (2021), agricultural soil management activities include application of synthetic and organic fertilizers, deposition of livestock manure, and growing N- fixing plants; therefore, it is difficult to quantify the contribution of livestock to this value. Yet, it is known that the fertilization of feed crops and deposition of manure on crops and pastures generate substantial amounts of N2O emissions (Steinfeld et al., 2006; Gerber et al., 2013). Nitrogen is an important macronutrient for plants that is used to improve crop yield; however, plants often do not uptake all the reactive N available in fertilizers and manure. This results in runoff of N, which eventually reacts with soil and produces N2O as a byproduct (EPA, 2020).

Globally, livestock production and feed crop production produce 65% of N2O emissions (EPA, 2020), although this number is difficult to discern in the US.

Nitrous oxide can be produced via microbial decomposition of organic nitrogen in livestock manure during management and storage (EPA, 2021; Gerber et al., 2013). The EPA (2021) estimated that manure management was the 4th largest contributor of anthropogenic N2O emissions, representing 4.3% of total emissions in 2019. These Figure 5. Nitrous oxide released via livestock manure. From De Klein et al., (2008) emissions are most likely to occur in unpaved beef and dairy cow drylots since these sites are susceptible to nitrification and denitrification processes, in which the byproduct is N2O (EPA, 2020) (Figure 5). In unpaved feedlots, NH4 that is not lost to volatilization will be absorbed by the soil and be oxidized to NO2 or NO3. The subsequent process of denitrification is dependent on on-site soil drainage – in poorly drained soils, anaerobic conditions needed for denitrification will be more likely, and therefore, will result in more N2O emissions (EPA, 2020). In 2019, N2O emissions from manure management and storage were approximately 20 MMT CO2-eq, representing an increase of 5.6 MMT CO2-eq since 1990, primarily due to the increase in specific livestock populations (EPA, 2021).

3.2.3. CO2 Carbon dioxide is considered the most prevailing GHG despite having the lowest radiative forcing, because the vast majority of GHG emissions are CO2, and it has a longer atmospheric lifetime

(EPA, 2020; IPCC, 2007). CO2 has a GWP of 1 CO2-eq over 100 years, and anthropogenic emissions will have effects that last thousands of years (EPA, 2021; IPCC, 2007). In the US, the livestock sector contributes to anthropogenic CO2 emissions primarily through fossil fuel consumption and land use change, although the relative contribution of the livestock sector is difficult to quantify.

The livestock sector in the US consumes fossil fuels for activities such as facilitating the Haber- Bosch process to fertilize feed crops and pastures, providing temperature regulation and ventilation in confined animal feeding operations (CAFOs), and for transportation of livestock and feed crops (Gerber et al., 2013; Pelletier et al., 2014; Steinfeld & Wassenaar, 2007). The Haber- Bosch process is an energy intensive artificial nitrogen fixation process that produces ammonia, which is primarily used for agricultural fertilizer (FAO, 2006; Steinfeld et al., 2006). In the US, approximately 4.7 million tons of synthetic fertilizer are used per year for feed crops and pasture lands (FAO, 2006). This equates to approximately 51% of synthetic fertilizer (primarily produced via Haber-Bosch) being used for livestock feed crops and livestock pasture, resulting in CO2 emissions of 11.7 MMT CO2-eq per year (FAO, 2006). Additionally, fossil fuels are used to regulate temperature and ventilation in CAFOs. In a study looking at broiler chicken sector life cycle assessments (LCAs), Pelletier et al., (2008) estimated that on-farm emissions and inputs were largely related to heating and ventilation of facilities and contributed approximately 226 kg CO2- eq per ton of broiler poultry produced. Facility energy use was also found to be significant in swine production systems in the Midwest; however, the contribution of GHGs from swine production are primarily due to manure management and production of feed (Pelletier et al.,

2010). Lastly, the livestock sector emits CO2 via the processing and transport of livestock animals. Global estimates indicate that processing and transport of dairy and beef cows emit 87.6 MMT

CO2-eq and 12.4 MMT CO2-eq, respectively (Gerber et al., 2013).

The conversion of forests and natural ecosystems to agricultural land and livestock pasture releases CO2 into the atmosphere and reduces the carbon sequestration potential of an ecosystem (Rojas-Downing et al., 2017). Globally, land use change is estimated to contribute 9.2% to the livestock sector’s overall emissions, with 6% attributed to pasture creation, and 3% attributed to feed crops (Gerber et al., 2013), although these values do not consider foregone sequestration potential and carbon storage potential. The IPCC (2013) reported that roughly 32%, or 180 GtC (gigaton carbon) of total GHGs (predominantly CO2) released into the atmosphere between 1750-2011 can be ascribed to land use change. Additionally, approximately one-third

18 of ice-free land on the planet is dedicated to animal agriculture, making the livestock sector a significant contributor to this figure (Rojas-Downing et al., 2017; Steinfeld et al., 2006). Steinfeld et al., (2006) estimated that livestock-related cultivated soils generated roughly 28 million tons of CO2 per year and livestock-induced desertification of pastureland generated roughly 100 million tons of CO2 per year on a global scale. As previously mentioned, livestock pastureland is the single largest land use in the US, and the production of livestock feed utilizes 40% of cropland (Bigelow & Borchers, 2017; Eshel et al., 2014), taking up land that could otherwise be reforested to sequester CO2.

3.3. Pollution The Environmental Protection Agency (EPA) indicated that at least 485,000 km of streams and more than 2 million hectares of lakes in the US did not meet water quality standards in relation to pollution (EPA, 2000). Hence, the widespread contribution of animal agriculture to aquatic ecosystem pollution has been a longstanding concern in the US (Burkholder et al., 2007). Animal agriculture is responsible for releasing pollutants such as nutrients, excess sediment, pharmaceuticals, and estrogenic compounds into nearby surface waters through pathways such as surface runoff, overflow of manure lagoons, direct disturbance or defecation, and atmospheric deposition (Burkholder et al., 2007; EPA, 2012; Hutchins et al., 2007; Kauffman et al., 1983; Mallin et al., 2015; Stone et al., 1995; West et al., 2011). It is important to note that in the studies found for this review, sedimentation is primarily associated with pastured or grazing animals, whereas nutrient, pharmaceutical, and estrogenic pollution are primarily associated with CAFOs. Since the 1950’s (chickens), and the 1970’s-1980’s (cows, pigs), most animals consumed in the US are raised in CAFOs rather than pastures (Burkholder et al., 2007). Of the 20,000 CAFOs in the US, only 8,000 facilities have National Pollutant Discharge Elimination System (NPDES) permits, which help address water pollution by regulating the amount and types of pollutants being discharged by point sources (EPA, 2012).

3.3.1. Nutrients Animal agriculture contributes large amounts of excretory nitrogen (N) and phosphorous (P) to aquatic ecosystems, primarily via overflow of waste lagoons, or runoff from recent applications of waste to crop fields (Burkholder et al., 2007). Waste generated by CAFOs is typically hosed through slats in the floor of the confinement building, where it is then drained and pumped into outdoor waste lagoons. From there, the waste either remains in the lagoon or is sprayed onto nearby fields to absorb excess nutrients (Mallin et al., 2015). Most studies relating to livestock waste and aquatic nutrient pollution focus on CAFOs; however, pastured animals contribute to increased nutrient concentrations via direct defecation in surface waters or runoff from pastures (James et al., 2007), and aquaculture animals contribute via waste flows of metabolic and food waste (Loch et al., 1996). Based on estimates from 2005, animal agriculture in the US produces 1.2 - 1.37 billion tons of manure per year, which is 3-20 times more solid waste than human sanitary waste production (EPA, 2005). The contribution of N and P from livestock systems causes a significant increase in eutrophication of aquatic systems, as well as increased ammonia levels that are toxic to aquatic species (Burkholder et al., 1997; Mallin, 2000; Mallin et al., 2015).

Myriad studies demonstrate the significant impact that nutrient pollution from CAFOs have on water quality. In a study conducted in a North Carolina watershed with swine and poultry CAFOs, Mallin et al. (2015) found that nitrate concentrations were consistently high throughout the watershed, reaching concentrations of >10mg N/L, and ammonium concentrations near swine waste spray-fields reaching up to 38 mg N/L. This study also found that phosphorous in animal waste was significantly correlated with high biochemical oxygen demand in samples (Mallin et al., 2015). Similarly, Westerman et al. (1995) reported that surface runoff from spray-fields that received swine waste at recommended concentrations had 3-6 mg NO3/L, while Stone et al. (1998) found 6-8 mg total inorganic N/L and 0.7-1.3 mg P/L in a stream bordering a spray-field. Therefore, nutrient losses from spray-fields receiving recommended concentrations of waste can surpass nutrient levels (100-200 µg inorganic P or N/L) known to support noxious algae blooms (Mallin, 2000). Sauer et al. (1999) measured nutrient runoff from test plots that contained chicken litter and found that, after a rain event, these sites transported 5.0%, 29.5%, and 21.9%

20 of the total nitrogen, ammonia, and reactive phosphorous that was applied. Additionally, studies conducted after a waste lagoon overflow or spill have demonstrated the substantial impacts these events have on aquatic ecosystems (Burkholder et al., 1997; Mallin, 2000). In 1995, the New River in North Carolina received 25 million gallons of liquid swine waste due to a waste lagoon rupture, which catalyzed algal blooms due to the nutrient load, and resulted in the pollution of approximately 22 river miles (35.4 km) (Burkholder et al., 1997). That same year, Mallin et al. (1997) documented the effects of a swine lagoon leak and a poultry lagoon breach that resulted in algal blooms in the Cape Fear River basin.

3.3.2. Sediment In a study conducted in 1982, excessive sedimentation was estimated to impact 46% of all US streams and rivers (Judy et al., 1984). Sediment has been identified as the most detrimental aquatic pollutant by some researchers, since it alters light attenuation, oxygen regimes, benthic habitats, and trophic food webs (dos Reis Oliveira et al., 2019; Henley et al., 2000; Ritchie, 1972). Animal agriculture significantly contributes to sedimentation, and literature pertaining to this topic primarily focuses on how overgrazing of riparian zones contributes to sedimentation, although CAFOs and aquaculture farms are contributors as well (Burkholder et al., 1997; Roberts et al., 2009). Moreover, feed crop agriculture also triggers significant sedimentation due to tilling and removal of riparian zones (Pimentel et al., 1995).

As previously mentioned, grazing cattle degrade riparian habitats via trampling and ingestion of riparian vegetation, which in turn promotes erosion of stream banks, especially during periods of precipitation (Braccia & Voshell, 2006). Vidon et al. (2007) found that cattle access to streams increased total suspended solid concentrations by 11 times, and turbidity by 13 times during summer and fall periods. In earlier studies, Winegar (1977) noted that a 3.5 mile (5.6 km) stretch of stream protected from grazing contained 48-79% less sediment compared to when it was previously grazed. Kauffman et al. (1983) found that during late season grazing periods, an average of 13.5 cm of streambank was lost, or eroded away, in grazed areas while 3.0 cm was lost in ungrazed areas. Similarly, other studies have found significantly higher total suspended

21 solids and lower streambank stability due to erosion in grazed areas compared to ungrazed areas (Grudzinski et al., 2018; Scrimgeour & Kendall, 2003). In addition to grazing cattle, swine waste spills have been found to significantly increase total suspended solids due to both swine waste and erosion (Burkholder et al., 1997). In North Carolina, a swine waste spill resulted in a 10-100- fold increase in suspended solids (SS) compared to unaffected reference sites, which equated to 71.7 mg SS/L (Burkholder et al., 1997). Although less studied, trout farms release fecal waste and food particulates in the form of organic suspended solids, which blanket streambeds and increase substrate embeddedness (Roberts et al., 2009). Animal agriculture further contributes to sedimentation of streams through its dependence on feed crop agriculture. In 1995, it was estimated that US crop agriculture alone created approximately 17 tons/ha/year of eroded soil, and 60% of this tonnage was deposited into waterways (Pimentel et al., 1995). It is important to note that, although this estimated tonnage considers all cropland in the US, a significant portion of it is dedicated to animal feed crops (Eshel et al., 2014).

3.3.3. Antibiotics Antibiotics are largely used as therapeutic treatments and growth inducers for livestock, as well as microbial treatments for aquaculture animals (Halling-Sorensen et al., 1998). Antibiotics released by livestock animals have been shown to provide selective pressure for the survival and spread of resistant organisms (Hooper, 2002). In 1998, one study showed that livestock consumed approximately 40% of the 22,680 tons of antibiotics produced by the US per year (Levy, 1998). It is estimated that 75% of antibiotics ingested by livestock end up in animal waste, eventually making their way into surface waters via runoff (Kolpin et al., 2002; West et al., 2011). In 2002, a survey of US surface waters unveiled residues of chlortetracycline, tetracycline, and oxytetracycline at maximum concentrations of 0.69, 0.11, and 0.34 µg/L, and found that water samples frequently contained mixtures of pharmaceuticals (Kolpin et al., 2002). West et al. (2011) examined water quality indicators in a watershed with CAFOS and observed that multi-drug- resistant bacteria were significantly more common at sites impacted by CAFOs.

3.3.4. Estrogenic compounds Animal agriculture waste flows are considered a source of estrogenic endocrine disrupting compounds (EDCs), often entering aquatic ecosystems via surface runoff (Campbell et al., 2006). Causal relationships between estrogenic EDCs in aquatic environments and CAFOs have been identified (Shrestha et al., 2012). Several studies have documented significant concentrations of estrogenic EDCs in streams and ponds adjacent to fields fertilized with chicken litter (Finlay- Moore et al., 2000; Nichols et al., 1997). More specifically, Nichols et al. (1997) recorded that EDC concentrations in runoff were dependent on rate of application, and documented EDC concentrations in chicken litter runoff as high as 1,280 ng/L. Total free estrogen concentrations (estrone, 17α-estradiol, 17β-estradiol, and estriol) have been found to be highest in swine lagoons (1,000-21,000 ng/L), followed by poultry lagoons (1,800-4,000 ng/L), then dairy lagoons (370-550 ng/L), and lastly, beef lagoons (22-24 ng/L) (Hutchins et al., 2007). For reference, EDC concentrations as low as 1 ng/L are known to negatively impact aquatic wildlife (Parrott & Blunt, 2004; Purdom et al., 1994).

3.4. Alien species Animal agriculture introduces alien species into aquatic ecosystems through the escape of farmed animals from aquaculture (Boersma et al., 2006; Pool, 2007; Thorstad et al., 2008). Fishes in aquaculture are selectively bred and genetically engineered to maximize their survival and productivity in aquaculture settings (Bostock et al., 2010; Thorstad et al., 2008). This genetic alteration results in significant genetic differences compared to wild fish populations; therefore, the breeding of wild and escaped genetically altered farmed species remains a concern (Bostock et al., 2010; Thorstad et al., 2008). Studies acknowledge that farmed fish escape from aquaculture pens, although these studies do not quantify the frequency of escapes (Pool, 2007; Thorstad et al., 2008). This section will remain specific to aquatic animals farmed for food and will not focus on animals used for sport fishing or bait.

In 2018, the US had 2,475 freshwater aquaculture farms (USDA, 2018). Channel catfish (Ictalurus punctatus) are the primary species produced in this industry, representing 287,132 tons, or 81%

23 of finfish produced (FAO, 2011). Channel catfish have a native range of approximately 20 states in the eastern-central portion of the US (Page & Burr, 1991), but are also produced outside of this range (FAO, 2011), and have been widely introduced throughout the country. The production of channel catfish occurs mostly in earthen ponds in the southeastern states of Louisiana, Arkansas, Mississippi, and Alabama, and is increasing in North and South Carolina (FAO, 2011, USDA, 1995). Other notable fish produced in freshwater aquaculture farms are rainbow trout (Oncorhynchus mykiss) and tilapia (Tilapia spp.). Rainbow trout are produced throughout the US, but are predominantly produced in northwestern Idaho (FAO, 2011), and are native to the Pacific northwest, western California, and western Idaho (Page & Burr, 1991). Throughout the decade of 1998-2008, production of rainbow trout was relatively stable, averaging approximately 24,000 tons per year (FAO, 2011). Tilapia species are grown throughout the US, although they are native to Africa and the Middle East (USDA, 1995). In 1994, 14,585 tons of tilapia were produced in the US (USDA, 1995). Another farmed fish of concern are Atlantic salmon (Salmo salar), which are anadromous fish that are grown in offshore net-pens, and swim to freshwater streams to mate and spawn when they escape (Thorstad et al., 2008). Atlantic salmon are primarily produced off the northeastern coast of the US, where the last wild populations remain, and are also produced on the west coast, where they are not native (FAO, 2011; NOAA, 2021).

3.5. Overexploitation Overexploitation of freshwater biodiversity for human consumption remains an understudied topic (Allan et al., 2005). Few studies quantify harvests of wild aquatic animals, and some studies claim that native ranges of certain fishes are unknown due to past overharvesting and subsequent collapse of stocks (Allan et al., 2005). Globally, total catch of freshwater fish was 8.7 million metric tons in 2002, with North America accounting for 2% of this value, equating to approximately 1.74 million metric tons (Allan et al., 2005). North America shows a declining trend in commercial freshwater fisheries; however, many of these deserted commercial fisheries have been replaced by recreational fisheries, which can add to total fish harvest but remain unreported (Cooke & Cowx, 2004).

Pacific salmon, Atlantic salmon, and Lake sturgeon (Acipenser fulvescens) populations are some of the few cases of overharvested freshwater species in the US that are cited in scientific literature (Allan et al., 2005; NOAA, 2021). Historically, Pacific salmon were overharvested to a degree that still affects present runs (Allan et al., 2005; Johnson et al., 2018). Commercial overharvesting of Chinook salmon in the Columbia River, in particular, have experienced 4 phases of overexploitation: (1) 1866-1888 is considered the beginning of the fishery, (2) 1889-1922 is considered the productive phase of harvesting, (3) 1923-1958 is the period of notable decline in Chinook salmon runs, (4) 1958-present agencies are now maintaining reduced productivity of runs (Johnson et al., 2018). Recently, NOAA (2018) designated five Pacific salmon stocks – two Chinook and three coho – as “overfished” and one Chinook stock as “subject to overfishing”. Harvest of summer-run Chinook salmon stock in the Upper Columbia River exceeded management plan recommendations by 14% in 2015 (NOAA, 2021). Moreover, wild Atlantic salmon in the US were historically native to nearly every coastal river northeast of the Hudson River in New York; however, overfishing reduced population sizes until the commercial fisheries were closed in 1948 (NOAA, 2021). The fishery is still closed, as wild Atlantic salmon in the US are listed as Endangered under the Endangered Species Act (ESA), yet overfishing remains a concern (NOAA, 2021). Wild Atlantic salmon are still captured due to fishing industries, either as bycatch or due to illegal poaching (NOAA, 2021). Lake sturgeon are another example of a freshwater species with documented cases of overexploitation (Allan et al., 2005). Beginning in the mid- 1800s, lake sturgeon were intensely harvested for their caviar and meat and were also intentionally overharvested because they damaged fishing gear intended for more valuable species (Allan et al., 2005; Boreman, 1997). Between 1879 and 1900, the Great Lakes commercial sturgeon fishery brought in an average of 1800 metric tons per year (USFWS, 2021). By the 1920s, lake sturgeons in the Great Lakes were depleted enough that the fishery closed (Allan et al., 2005).

4. State

Each of the five pressures have negative consequences on aquatic biodiversity. The following sections will look at the state of aquatic biodiversity as a result of each pressure. Since some pressures are interrelated, a multistress section is also included.

4.1. Habitat alteration The loss of riparian zones due to grazing significantly diminishes the availability of suitable habitat for aquatic biodiversity by decreasing shade, compacting soil via trampling, and decreasing allochthonous inputs (Batchelor et al., 2014; Goss & Roper, 2018; Kauffman et al., 1983; Nussle et al., 2017). Salmonids are especially vulnerable to these effects, as they are temperature- sensitive, cold-water species. Goss & Roper (2018) evaluated 153 stream reaches in the Columbia Basin, and noted that improved stream conditions for salmonids – salmon, trout, and char – corresponded to decreased livestock disturbance along riparian zones. A review conducted by Platts (1991) reported that 15 of 19 studies showed that salmonid populations declined in the presence of livestock, whereas Behnke (1978) revealed that golden trout (Oncorhynchus mykiss aguabonita) biomass was 3-4 times greater in ungrazed sections of streams when compared to grazed sections. Although myriad studies focus on the decline of salmonids in relation to livestock grazing in riparian zones (Armour et al., 1994; Elmore & Beschta, 1987; Kauffman & Krueger, 1984; Moyle et al., 2016; Van Velson, 1979; Zoellick, 2004), there are also plenty of studies focusing on aquatic invertebrates. Frest (2002) identifies livestock grazing as a primary factor in the extirpation of freshwater snail populations via direct trampling and soil compaction. In Virginia (VA), the Piedmont pondsnail (Stagnicola neopalustris), which was known to be endemic to a single pond, is understood to be extinct due to cattle grazing (Herrig & Shute 2002). Moreover, macroinvertebrate communities tend to shift from intolerant taxa (e.g., Ephemeroptera, Plectoptera, Trichoptera) to more tolerant taxa (e.g., chironomids) in the presence of cattle grazing and loss of riparian habitat, partially due to the decrease in food quality and loss of habitat (Belsky et al., 1999; Jackson et al., 2015; Rinne, 1988; Tait, et al., 1994). In various studies, macroinvertebrate richness declined in study areas with livestock grazing compared to ungrazed areas (Herbst et al., 2012; McIver & McInnis, 2007). In a study conducted

26 on montane streams in California, Herbst et al., (2012) found that all macroinvertebrate richness metrics were significantly lowered in grazed areas compared to ungrazed areas, and that macroinvertebrate richness metrics significantly increased after a 4-year period of no grazing. Several authors mention that studies conducted on macroinvertebrates assemblages in relation to livestock grazing were difficult to draw direct conclusions from, since basin-wide effects confounded the effects of livestock grazing (McIver & McInnis, 2007; Weigel et al., 2000).

Channelization homogenizes aquatic habitats that are important for biodiversity, such as riffles and pools, by altering width and depth of streams (Gorman & Karr, 1978; Smiley & Gillespie 2010). Aquatic biodiversity frequently increases as habitat complexity increases, with depth and velocity being crucial factors (DosSantos, 1985; Felley & Felley, 1987; Gorman & Karr, 1978; Pusey et al., 1995; Schlosser, 1982). Therefore, channelization has been found to decrease fish abundance (Hansen, 1971; Lau et al., 2006) and decrease fish community quality by increasing tolerant taxa and decreasing environmentally sensitive taxa (Etnier, 1972; Gorman & Karr 1978; Scarnecchia 1988; Sullivan et al., 2004). Channelization also decreases fish spawning success by increasing flow velocity due to additional runoff from crop fields (Carline & Klosiewski, 1985; Lau et al., 2006). In a study conducted by Lau et al., (2006), 8 fish species that are associated with pool habitats – green sunfish, common carp, black redhorse, golden redhorse, bluegill, smallmouth bass, spotted sucker, and white sucker fish – were all lost from channelized study areas. Channelization has also been found to alter benthic habitat important for freshwater mussels, potentially increasing the abundance of soft-substrate mussels, although there are few studies on this topic (Watters, 1999).

Aside from riparian zones and habitat heterogeneity, depletion of river flows for cattle feed crops has been found to be another crucial aspect of aquatic biodiversity loss (Reed & Czech, 2005; Richter et al., 1997; Richter et al., 2020). Richter et al. (1997) showed that 62% of sub-watersheds in the western US contain at least one species endangered by flow depletion, including two-thirds of all native fish in the Colorado River Basin. In a later study, Richter et al., (2020) estimated that 690 out of 1000 (70%) instances of increased risk of extinction for western fish species due to

27 flow depletion were caused by cattle feed crop irrigation. Richter et al., (2020) is perhaps the first study to link beef consumption to the depletion of fish species in specific river systems, despite the large water footprint of beef.

4.2. Climate change

4.2.1. Effects of increased in-water temperatures on aquatic biodiversity Animal product production significantly contributes to climate change by increasing GHG’s such as CH4, N2O, and CO2 (EPA, 2020; EPA, 2021; Gerber et al., 2013; IPCC, 2007; IPCC, 2013; Steinfeld et al., 2006). Studies have found that climate change results in increased air temperature, and therefore, increased in-water temperatures in freshwater systems (Mohseni & Stefan 1999; Webb et al., 2008). It is predicted that future increases in water temperature will heavily impact the physiological processes of aquatic biodiversity by altering local resource requirements, habitat tolerances, and skewing important temperature-related cues for the timing to life-history events (Whitney et al., 2016). The primary challenge in predicting the impacts of climate change on aquatic biodiversity is characterizing the physiological sensitivity of populations to changes in local conditions (Knouft & Ficklin, 2017). For example, in a study conducted on populations of bluntnose minnows (Pimephales notatus), Beachum (2014) recorded a gradient of variation in metabolic sensitivity that followed identical, positively correlated increases in water temperature and demonstrated that populations from warmer regions were more sensitive to changes in water temperature than populations from cooler regions. In contrast, stream macroinvertebrate populations from higher elevation (cooler) regions have been documented to be more sensitive, or have higher turnover rates, than populations from lower elevation (warmer) regions (Bush et al., 2012).

Increased in-water temperatures effect freshwater fish species and populations via phenological shifts, altered demographics, and distributional shifts. Studies conducted on lotic habitats have found that the combination of warmer winters, earlier spring warming, and warmer summers have resulted in earlier temperature-dependent life events in fish, such as migration and

28 spawning (Kovach et al., 2013; Quinn & Adams, 1996), although responses among fish are varied. For example, Peer & Miller (2014) found that at lower latitudes, striped bass (Morone saxatilis) responded to earlier spring warming by displaying earlier spawning migrations. At mid-latitudes, Atlantic salmon (Salmo salar), American shad (A. sapidissima), and sockeye salmon (Oncorhynchus nerka) responded to earlier spring warming and overall increases in spring and summer temperatures by beginning their spring migrations early (Crozier et al. 2011; Juanes et al., 2004; Otero et al., 2014; Quinn & Adams 1996). At higher latitudes, numerous Pacific salmon (Oncorhynchus spp.) populations have responded to warmer spring temperatures by migrating to the ocean earlier than normal (Kovach et al., 2013). Climate-induced increases in in-water temperatures have altered fish population demographics in aquatic ecosystems by decreasing recruitment, decreasing abundance, and in some cases increasing growth of juveniles (Bassar et al., 2016; Eby et al., 2014; Kovach et al., 2014). In a study conducted over the span of 17-20 years in the Rocky Mountains, the authors found that bull trout (Salvelinus confluentus) abandoned study sites more frequently than they colonized them, leading to a decrease in the number of occupied sites (Eby et al., 2014). Abandonment probabilities were significantly higher at lower elevations, supporting previous climatic models that predicted that bull trout would migrate to higher elevations in the face of increasing in-stream temperatures due to climate change (Eby et al., 2014). In another study conducted on eastern brook trout (Salvelinus fontinalis) in western Massachusetts, the authors discovered that the primary cause of decreased survival rates of juveniles was increasing mean stream temperatures (Bassar et al., 2016). Studies conducted on lotic habitats must consider confounding anthropogenic stressors; therefore, it is difficult to observe changes in thermal regimes due to climate change (Lynch et al., 2016). Yet, the consistent trend across species and populations strongly indicates that a shared environmental driver (climate change) is responsible (Lynch et al., 2016).

Increased in-water temperatures also impact aquatic macroinvertebrates populations; however, there are far fewer studies on how climate change effects macroinvertebrates in the US, compared to fish. Several studies have documented that the development of macroinvertebrate larvae is closely tied to temperature and is expected to be negatively affected by climate change

(Collins & McIntyre, 2017; Jordan et al., 2016). River odonates (damselflies and dragonflies) are indicators of aquatic health and are vulnerable to the effects of climate change because they have species-specific habitat requirements. Species distribution models (SDMs) demonstrate that climate change will decrease the amount of suitable habitat reaches for odonates by 45-99% by 2080 (Collins & McIntyre, 2017). In a study conducted on an endemic stonefly (Lednia tumana) in Montana, Jordan et al. (2016) found that genetic diversity of populations had significantly declined over 10 years, most likely due to glacial loss in response to climate change.

Lentic habitats, such as lakes and ponds, are directly affected by increased temperatures due to climate change, making these habitats excellent sentinels for climate change monitoring. Multiple studies have projected that increased temperatures due to climate change would prolong the stratification of lakes by approximately 2 months, and in some cases, lakes could remain stratified for the entire year (Blumberg & DiToro, 1990; McCormick, 1990). Alteration of mixing regimes in lakes could affect nutrient cycling and deplete dissolved oxygen levels (Blumberg & DiToro, 1990; McCormick, 1990), negatively impacting biodiversity (Jacobson et al., 2012). Jacobson et al. (2012) examined 634 lakes in Minnesota and documented a significant decline in stenotherm cisco (Coregonus artedi), with the timing of declines strongly corresponding to increased warming and longer periods of summer stratification, likely due to climate change. Increased temperatures in lakes have also resulted in earlier spawning by yellow perch (Perca flavescens) in Lake Erie (Lyons et al., 2015), and walleye (Sander vitreus) in Minnesota lakes (Schneider et al. 2010).

4.2.1. Effects of altered precipitation on aquatic biodiversity Increases in extreme flow events in the past century, such as flooding and drought, have been linked to climate change, and are projected to continue (Milly et al., 2002). Projections show that the southeastern US is likely to experience increased overall precipitation, with the exclusion of the northern, interior portion of the region. Precipitation in this region is also expected to occur as more intense, clustered storms; however, evapotranspiration is likely to exceed precipitation, and therefore, decrease runoff (Mulholland et al., 1997). Overall, increases in storm intensity and

30 clustering will result in greater peak flows and lower baseflows (Mulholland et al., 1997). In general, studies show that wet regions are projected to get more precipitation, and arid regions are projected to receive less precipitation (Dore, 2005). Along with aquatic temperature increases, altered flow regimes are one of the largest threats of climate change on aquatic biodiversity (Collins & McIntyre, 2017; Lynch et al., 2016).

Aquatic species and populations have responded to altered flow regimes as a result of climate change via reductions in recruitment and population sizes, and declining survival rates (Bassar et al., 2016; Collins & McIntyre, 2017; Ward et al., 2015). In an analysis of 21 chinook salmon (Oncorhynchus tshawytscha) populations, Ward et al. (2015) projected that increases in flow variability over the last 60 years occurred in more than half of studied rivers in the Pacific northwest and would be a leading cause of chinook salmon population declines. Similarly, Bassar et al. (2016) found that brook trout populations in Massachusetts decreased in size due to altered flows, although increased in-water temperatures were the main driver. Moreover, increased precipitation and flooding are predicted to negatively impact odonates in the northeastern US by decreasing survival and emergence rates (Collins & McIntyre, 2017; Thompson, 1990).

4.3. Pollution Nutrient pollution from animal agriculture degrades aquatic habitats by contributing to eutrophication of ecosystems, as well as increased ammonia levels that are often toxic to aquatic wildlife (Burkholder et al., 1997; Mallin, 2000; Mallin et al., 2015). Freshwater eutrophication due to animal agriculture has been found to promote noxious algal blooms, such as cyanobacteria, and can even result in long-term shifts in phytoplankton community structure from advantageous species, such as diatoms, to noxious species (Burkholder et al., 2007). Some algal blooms caused by animal agriculture contain harmful organisms, such as Pfiesteria piscicida, which has been shown to be a primary agent of major fish kills in Maryland and Virginia tributaries of the Chesapeake Bay (EPA, 1997). The subsequent decomposition of excessive algae lowers dissolved oxygen (DO) concentrations for aquatic organisms, further altering community structures (Burkholder et al., 2007). In Blue Ridge Mountain streams impacted by nutrient runoff from cattle

31 farms, lowered DO concentrations were found to promote the presence of macroinvertebrate taxa that are tolerant of low DO, while macroinvertebrate taxa that are oxygen-sensitive became absent (Braccia & Voshell, 2006). Burkholder et al. (1997) noted that two days after a swine manure spill in North Carolina, a 29 km stretch of river became anoxic and contained high levels of ammonia, which resulted in the death of approximately 4,000 freshwater fish.

Ammonia from animal agriculture runoff has been associated with fish, macroinvertebrate, and

+ mussel kills (O’Hearn & McAteer, 2016). Ammonia is toxic to fish because increased NH4 displaces K+, therefore depolarizing neurons and causing a cascade of chemical reactions in the central nervous system that result in convulsions, coma, and death (Randall & Tsui, 2002). A study conducted on fish kills in Missouri found that a 16 mile (25.7 km) stretch of river contained levels of ammonia that were toxic to aquatic life for 7 days, resulting in a kill of at least 17,474 fish, 23,636 benthic invertebrates (including mussels and snails), and 26 tadpoles (O’Hearn & McAteer, 2016). The suspected source of pollution was an egg farm that was releasing chicken wastewater and manure (O’Hearn & McAteer, 2016). Un-ionized ammonia has been recognized as a significant cause of freshwater mussel declines, yet studies do not explicitly link this issue with animal agriculture (Newton, 2004; Strayer & Malcom, 2012).

Sedimentation due to animal agriculture alters light attenuation and benthic habitats in aquatic ecosystems, triggering deleterious effects on aquatic wildlife (Braccia & Voshell, 2006; Henley et al., 2000; Richards & Host, 1993). Increased suspended sediment concentrations due to cattle grazing limit light penetration and therefore, phytoplankton production (Henley et al., 2000). As suspended sediment concentrations increase, primary production in aquatic ecosystems decreases, decreasing the amount of food available for herbivorous fish and macroinvertebrates, as well as zooplankton (Henley et al., 2000). McCabe & O’Brien (1983) found that turbidity levels as low as 6 NTU decreased the feeding efficiency of Daphnia pulex by 25%, demonstrating that zooplankton can suffer population declines with low sediment contributions. Excessive sedimentation due to animal agriculture also alters benthic habitats by contributing fine sediment and silt to interstitial spaces between cobble stones in stream beds, thereby increasing

32 the frequency of sand and silt patches (Braccia & Voshell, 2006). These habitat alterations alter macroinvertebrate communities and displace macroinvertebrate taxa that utilize interstitial spaces for crawling (Lenat et al., 1981; Wood & Armitage, 1997). In a study conducted on streams adjacent to feed crops in Michigan, Richards & Host (1993) recorded that macroinvertebrate community structure in agriculture-intensive streams differed the most from other sites, primarily due to the addition of fine sediments. In a study conducted in 1971, Stansbery recorded that sedimentation linked to agricultural land use intensity in the southeastern US was shown to trigger declines in 40 – 50% of freshwater molluscan fauna.

Antibiotics used in animal agriculture and aquaculture that enter aquatic ecosystems have been shown to affect aquatic biodiversity (Brain et al., 2005; West et al., 2011; Wollenberger & Kusk, 2000), although studies on this topic in the US are sparse. As previously mentioned, antibiotics entering freshwater ecosystems select for antibiotic-resistant bacteria, altering the aquatic bacterial community (Hooper, 2002; West et al., 2011). West et al. (2011) noted antibiotic- resistant bacteria at all study sites downstream of CAFOs, and laboratory assays showed conjugative transfer of resistance was frequent in these bacterial communities. The study also found that the proportion of multi-drug resistant bacteria was 41.6% at sites receiving manure downstream of CAFOs, and that these sites reflected site-specific antibiotic-resistance patterns (West et al., 2011). A toxicity test conducted on freshwater crustacean Daphnia magna showed that 50% of populations displayed decreased reproductive output at concentrations of 46.2 mg/L and 44.8 mg/L for oxytetracycline and tetracycline, respectively (Wollenberger et al., 2000). Brain et al. (2005) demonstrated that the macrophyte Myriophyllum sibiricum had a significant concentration-response relationship with a mixture of tetracyclines. Dry mass of M. sibiricum was 69, 47, 30, and 7% of controls at respective tetracycline concentrations of 0.080, 0.218, 0.668, and 2.289 µmol/L (Brain et al., 2005). It is important to note that the aforementioned toxicity studies were conducted in laboratory settings, rather than observed in agricultural settings.

Manure and urine from animal agriculture are contributors of estrogenic EDCs in aquatic environments (Campbell et al., 2006). Estrogenic EDCs have recently received widespread

33 attention in aquatic ecology, primarily since male fish exposed to estrogenic EDCs are known to produce vitellogenin, an egg yolk precursor protein, which produce eggs in the testes of males and results in intersex (Orlando et al., 2004; Yonkos et al., 2010). In one study, Yonkos et al. (2010) exposed mature male Pimephales promelas individuals to samples that contained runoff from a poultry-litter amended field for 21 days and documented significant vitellogenin production in individuals. Similarly, Orlando et al. (2004) collected wild Pimephales promelas individuals exposed to CAFO feedlot effluent and found that male fish had lower testicular testosterone synthesis, altered head morphometrics, and smaller testes sizes. These estrogenic compounds have been shown to negatively impact fish species at exposure concentrations as low as 1.0 ng/L (Parrott & Blunt, 2004; Purdom et. al., 1994). Moreover, it is also known that EDC concentrations in chicken litter runoff can be as high as 1,280 ng/L (Nichols et al., 1997) and can be present in agricultural manures and wastewaters at concentrations of 40–7000 ng/L (Tremblay et al., 2018); therefore, more research should be done on environmental concentrations of EDCs in agricultural runoff and the associated impacts on aquatic wildlife.

4.4. Alien species Alien species introduced into aquatic ecosystems through aquaculture have been shown to alter genetic diversity of wild populations, compete with wild aquatic animals for resources, and feed on native species (Boersma et al., 2006; Fleming et al., 2000; Hill, 2008; USFWS, 2002). The primary freshwater species that are farmed in the US for food – channel catfish, rainbow trout, tilapia, and Atlantic salmon – have all been shown to exhibit negative effects on native freshwater populations (Boersma et al., 2006; Fleming et al., 2000; Hill, 2008; USFWS, 2002). Channel catfish introduced via aquaculture have established populations outside of their native ranges, as seen in the western US (Pool, 2007). Channel catfish are successful invaders in the western US due to the plasticity of their diets, as well as their natural defenses against predators (e.g., spines on dorsal and pectoral fins) (Boersma et al., 2006; Bosher et al., 2006). Boersma et al. (2006) observed the gut contents of alien channel catfish in Columbia River reservoirs and found that 50-100% of their gut contents were native salmon prey, and that this value was dependent on the size class of the catfish. In Arizona, Marsh & Brooks (1989) noted that channel catfish in the

Colorado River displayed high predation pressure on native razorback suckers (Xyrauchen texanus) that were being reintroduced to the basin, resulting in 900 deaths of the razorback suckers along a 2.5 km study area, although this number considers non-native flathead catfish predation as well. Escaped rainbow trout are a concern in areas where closely related native species exist, such as the rare cutthroat trout (Oncorhynchus clarkii) (Hill, 2008). Since rainbow trout are genetically altered in aquaculture, if genes from farmed fish are passed into the genome of native fish, it is possible that rare native stocks could be inundated with new genetic material and become extinct (Hill, 2008). The invasion of farmed tilapia species is a large concern in the US, since tilapias are non-native to the US, and have substantial impacts on aquatic biodiversity (Peterson et al., 2002; USFWS, 2002). After tilapia escaped from an aquaculture facility in Mississippi in 2002, Peterson et al. (2002) found that tilapia could feed at any trophic level, and ingested native microcrustaceans and macroinvertebrates. The study also found that native bluegill and largemouth bass declined in abundance due to the aggression that tilapia exhibit during mating, as well as a decrease in habitat related to tilapia spawning (Peterson et al., 2002). USFWS (2002) proposed to eradicate tilapia from the Virgin River in Nevada due to predation on native endangered woundfin (Plagopterus argentissimus) and endemic Virgin River chub (Gila seminuda). Similar to rainbow trout, farmed Atlantic salmon are an environmental concern due to their genetic engineering. Although farmed Atlantic salmon are not an alien species in the northeastern US, they negatively impact the genetic diversity of endangered wild populations (Fleming et al., 2000). Hybrid juveniles, or offspring of one farmed and one wild Atlantic salmon, exhibit an intermediate trait expression (Fleming et al., 2000). This is especially concerning because farmed Atlantic salmon exhibit lower fitness than their wild counterparts, making the hybrid offspring less fit in the wild (Fleming et al., 2000). Farmed Atlantic salmon also compete with wild populations for resources, and often have a competitive advantage due to their selective breeding for growth (Fleming et al., 2000). In an experimental release study, Fleming et al. (2000) found that the native juvenile Atlantic salmon population decreased by more than 30% when farm and hybrid individuals were present. Atlantic salmon that escape on the west coast, outside of their native range, have negative impacts on native fish species in streams (NOAA, 2021). In Pacific streams, Atlantic salmon compete with rare steelhead trout for food and habitat,

35 and often have a competitive advantage due to their increased size at a young age (Volpe et al., 2000).

4.5. Overexploitation In the US, it is known that many freshwater species are harvested for food, yet few studies point to overharvest as a significant factor for biodiversity loss of certain species. Pacific salmon, Atlantic salmon, and Lake sturgeon have all been shown to suffer from reduced numbers due to human consumption (Allan et al., 2005; Johnson et al., 2018; NOAA, 2021; USFWS, 2021). The historical overharvesting of Pacific salmon in the late 1800s is still a significant factor in their decreased numbers today (Allan et al., 2005; Johnson et al., 2018). Currently, Pacific salmon are stocked from hatcheries to make up for their decreased numbers from overharvesting, as can be seen on the Columbia River. By 1995, as much as 80% of Chinook salmon seen on the Columbia River were hatchery-stock salmon (Johnson et al., 2018). Johnson et al. (2018) analyzed the genetic diversity of contemporary Chinook salmon, compared to ancient salmon, and found that genetic diversity was significantly reduced in contemporary stocks, as measured in both direct loss of mitochondrial haplotypes as well as declines in haplotype and nucleotide diversity. The authors credited the loss of contemporary genetic diversity to the “4 H’s”- two of which are harvest and hatcheries (Johnson et al., 2018). Wild Atlantic salmon have also suffered population declines due to past overharvesting, the effects of which can still be seen in the overall lower number of individuals (NOAA, 2021). Currently, the last wild populations of US Atlantic salmon can only be found in eight rivers in Maine (NOAA, 2021). To rebuild population numbers, hatcheries were created to produce wild Atlantic salmon. In 1985, hatcheries produced 5,000 adult salmon that were released into the wild; however, annual returns to the US were generally less than 1,000 adults for numerous reasons, making this species a priority for NOAA to prevent its extinction (NOAA, 2021). In addition, Lake sturgeon populations continue to decline due to overharvesting, and the species is currently listed as threatened in 19 out of the 20 states it inhabits (USFWS, 2021). Lake sturgeon are particularly susceptible to populations decline brought on by overfishing due to their slow reproductive cycle (USFWS, 2021). Most sturgeons are harvested before they reach 20 years of age, and therefore, have never bred (USFWS, 2021).

Although there are studies that point to overharvesting as a significant contributor of biodiversity loss for Pacific salmon, Atlantic salmon, and Lake sturgeon, no studies were found on harvesting effects on US crayfish populations or turtle populations, despite the demand that exists for their meat and eggs.

4.6. Multi-stress The pressures of habitat loss, climate change, pollution, biological invasions, and overexploitation simplify aquatic food webs by focusing energy flow along less pathways, thereby imperiling long- term community perseverance (Heleno et al., 2020). Many aquatic communities are subject to a combination of these pressures due to animal agriculture, resulting in simultaneous stressors for aquatic biodiversity (Lynch et al., 2016; Richter et al., 1997). There are myriad combinations of stressors that animal agriculture can pose for aquatic biodiversity, and often these stressors have synergistic effects (Lynch et al., 2016; Richter et al., 1997). In a study conducted on intensive agriculture sites that were recently converted from woodlands and prairies in Iowa, Poole & Downing (2004) found that more than half of the 118 stream reaches samples lost >75% of freshwater mussel species. The greatest mussel declines were significantly correlated with loss of riparian canopy and siltation associated with intensive agriculture (Poole & Downing, 2004). Other studies conducted on freshwater mussels have found that populations are also threatened by a combination of sedimentation and destabilized stream bottoms (i.e., habitat alteration) from feed crop agriculture runoff (Neves et al., 1997). Anthropogenic flow reductions caused by water extraction for cattle can cause a variety of multi-stress scenarios for freshwater fauna (Richter et al., 2016). According to Richter et al. (2016), anthropogenic flow reductions can concentrate existing nutrient and chemical pollution, resulting in eutrophication and reduced oxygen levels for aquatic fauna. Flow reductions can also shift competitive advantage to aquatic alien or invasive species, which has already been recognized as an issue in western states (Richter et al., 2016). The combined stressors of channelization, water extraction, and climate change have resulted in reduced fish habitat size, complexity, and lateral connectivity with floodplain habitats within the Rio Grande, ultimately lowering biotic richness (Lynch et al., 2016). These are a few

37 examples of the documented multi-stress scenarios that can occur in aquatic ecosystems due to animal agriculture.

5. Potential responses

Animal product consumption impacts aquatic biodiversity across the US; therefore, it is important to investigate potential responses that can be taken to mitigate this pressing issue. This section will look at a few approaches to mitigating the pressures of animal product consumption on aquatic ecosystems. It remains clear in the literature that animal product consumption and consumer demand have the highest potential for improving environmental conditions that impact biodiversity (Gerber et al., 2013; Machovina et al., 2015; Poore & Nemecek, 2018), yet production also offers room for improvements, hence this paper addresses both.

5.1. Production The aquatic impacts of animal product consumption may be partially mitigated through production or supply-side interventions. One potential response is the wide-spread implementation of Best Management Practices (BMPs) in grazing systems (USDA-NRCS, 2021). BMPs recommended by the National Resources Conservation Service (NRCS) include livestock exclusion fencing and riparian buffer strips (USDA-NRCS, 2021). Livestock exclusion fencing aims to keep cattle out of waterways, and therefore eliminates streams as a water source for cows. In a study conducted on the effectiveness of livestock exclusion fences, Owens et al. (1996) noted a 40% decrease in soil losses and a 57% decrease in annual flow weighted averages for sediment concentrations in heavily grazed streams in Ohio. Yet, Sarr (2002) noted that livestock exclusion fencing presents mixed results when considering geomorphological stream changes, with some studies citing positive geomorphic changes (e.g., channel narrowing), and others documenting no recovery after fence installation. There is also evidence that riparian vegetation recovers after livestock exclusion fencing has been implemented; however, the rate of riparian recovery can vary considerably between sites, with some stream sites taking decades to significantly recover after grazing (Kauffman & Krueger, 1984). Another BMP recommendation made by NRCS are

38 restoring or maintaining riparian buffer strips on cattle farms (USDA-NRCS, 2021). Dosskey (1998) found that riparian buffer strips consisting primarily of trees and shrubs were “highly effective” at stabilizing stream banks and “moderately effective” at filtering dissolved orthophosphates and nitrate-nitrogen. It is important to note that these responses may present economic constraints for farmers; therefore, their implementation may be dependent on economic conditions or available funding mechanisms.

Increasing agricultural input efficiency in animal production systems is another potential supply- side response to mitigating pressures on aquatic biodiversity. Agricultural input efficiency refers to the amount of food produced per unit of agricultural input (Clark & Tilman, 2017). Higher input efficiency is often associated with decreased environmental impacts, although the environmental benefits vary across systems (Clark & Tilman, 2017). Several ways to increase input efficiency in livestock systems is to increase feed efficiency, utilize agricultural waste, and shift from grass-fed or pasture systems to grain-fed or CAFO systems for cows (Clark & Tilman, 2017). One significant way to increase feed efficiency for livestock would be to supplement amino acids in diets (Clark & Tilman, 2017). Ogino et al. (2013) noted that pigs fed diets supplemented with amino acids utilized less feed, emitted 5% less GHGs, and had 28% less eutrophication potential compared with pigs fed diets without supplements. Similar results were found in chicken, beef cow, and dairy cow systems (Robertson & Vitousek, 2009). Utilizing agricultural waste products as animals feed is also an effective way to increase input efficiency. According to zu Ermgassen et al. (2016), this response could reduce the environmental impacts of livestock production by 20%, without reducing food quality for livestock animals. Lastly, shifting from grass-fed beef cows to grain-fed beef cows also significantly increases agricultural input efficiency (Clark & Tilman, 2017). Grass-fed beef has higher land-use requirements, and on average 19% higher GHGs per unit of food compared to grain-fed beef (Clark & Tilman, 2017). These inefficiencies of grass-fed beef stem from the low nutritional value of grass, silage, and fodder, as well as the grass-fed cows’ 6-12 month longer life span (Clark & Tilman, 2017). It is important to note, however, that increases in agricultural input efficiencies may make livestock agricultural activities more attractive and profitable, thereby increasing animal product consumption (Poore

& Nemecek, 2018). In addition, shifting to grain-fed beef may pose ethical concerns as animals face greater confinement.

5.2. Consumption Producers play an important role in responding to the pressures of animal agriculture, yet their ability to reduce environmental impacts is limited. These limitations stem from the fact that certain food products are inherently more environmentally detrimental than other nutritionally equivalent products, regardless of their production method (Poore & Nemecek, 2018). As previously mentioned, animal products are inherently more inefficient than plant-based products due to the energy transfer that occurs between trophic levels when converting plant calories into animal calories (Cassidy et al., 2013; Swain et al., 2018). Hence, a shift towards plant-based diets, along with a decrease in consumer demand for animal products would significantly reduce the environmental pressures of our current food systems (Aleksandrowicz et al., 2016; Harwatt et al., 2017; Poore & Nemecek, 2018). To a lesser extent, shifting diets from energy-intensive animal products to lower-impact animal products also creates environmental benefits (Aleksandrowicz et al., 2016; Poore & Nemecek, 2018).

A shift in human diets is the most impactful response to the pressures that animal agriculture place on aquatic biodiversity (Machovina et al., 2015; Poore & Nemecek, 2018). In a study comparing categories of sustainable diets, Aleksandrowicz et al. (2016) found that the largest environmental benefits in GHG’s, land use, and water use occurred in diets that most reduced the consumption of animal products. Pescatarian diets ranked second and third place for GHG’s and land use benefits, while vegetarian diets had the largest benefits in terms of water use (Aleksandrowicz et al., 2016). Plant-based diets (excluding all animal products), however, ranked first in land use and GHG benefits, with median reductions of 45% and 51%, respectively (Figure 6A,B) (Aleksandrowicz et al., 2016). Diets that replaced ruminants with monogastric animals resulted in some environmental benefits yet were significantly less effective than plant-based diets (Aleksandrowicz et al., 2016). Similarly, Harwatt et al. (2017) noted that substituting beef for beans - in terms of calories and protein - would help the US achieve 46-74% of the GHG

40 reductions needed to meet the 2020 emissions target, as well as free up 42% of US cropland (692,918 km2) and reverse the impacts of overgrazing. According to Peters et al. (2016), 735 million Americans could be sustained on a healthy plant-based diet using current croplands, compared with 402 million Americans on current diets, demonstrating the lower resource demand of plant-based foods. Animal-excluding diets at the global level could reduce agricultural land-use by 76% (Poore & Nemecek, 2018), which is equivalent to an area more than three times the size of the United States, including Alaska, Hawaii, and Puerto Rico. In a newer study, Hayek et al. (2021) projected that a global adoption of plant-based diets would result in 358-743 GtCO2 removal via carbon sequestrations through restoration, which is roughly equivalent to the past 9 to 16 years of fossil fuel emissions. In addition, this global shift to plant-based diets would allow for reforestation of pasture lands, sequestering considerable amounts of carbon out of the atmosphere. One study found that restoring artificial pastures to 100% carbon storage potential (i.e., rewilding and reforesting) could sequester roughly 108 PgC (Erb et al., 2018). For reference, 4 PgC accumulated in the atmosphere annually from 2000–2009 (IPCC, 2013). This combination of decreasing land use and decreasing GHG’s will alleviate pressures placed on aquatic biodiversity, with positive cascading effects. For example, if livestock animals are not consumed or consumed to a much lesser extent, there will likely be less contributions to stream bank slumping, riparian degradation, nutrient pollution, sedimentation, antibiotic pollution, and estrogenic pollution. Although the literature focuses less on aquatic animal consumption, it is likely that a reduction or elimination of aquatic animal consumption will lessen alien species introduced by aquaculture, as well as overexploitation.

A. GHG emissions

B. Land use

Figure 6. (A) Relative differences in GHG emissions (kg CO2eq/capita/year) between current average diets and sustainable dietary patterns. (B) Relative differences in land use (m2/capita/year) between current average diets and sustainable dietary patterns. Note: n = number of studies, mdn = median. From Aleksandrowicz et al. (2016).

5. Conclusions

This review aimed to link scientific articles that studied the contributions of animal agriculture to the five primary pressures of aquatic biodiversity loss, and articles that studied the state of aquatic biodiversity as a result of these pressures. The information compiled in this literature review demonstrates that there is a strong link between animal product consumption and aquatic biodiversity loss, and that animal agriculture is likely a primary driver of aquatic biodiversity loss. In the US, livestock grazing land is the single largest land use type and 40% of all grown crops are fed to livestock, leading to aquatic riparian zone degradation, channelization, and water extraction that impacts freshwater fauna. Animal agriculture also results in significant contributions of CH4, N2O, and CO2, contributing approximately 14.5% to global anthropogenic GHG emissions. In turn, climate change has been shown to increase in-water temperatures and alter precipitation patterns, which have deleterious impacts on freshwater fauna. Moreover, animal agriculture generates excess nutrients, excess sediment, antibiotics, and estrogenic compounds, which have all been shown to runoff into adjacent waterways and impact aquatic wildlife. Aquaculture is responsible for the escape of alien fish species in freshwater ecosystems, which affect native species by altering genomes via interbreeding, competing for resources, and consuming native individuals. Lastly, direct consumption of wild freshwater fauna by humans has resulted in decreased population sizes; however, this topic needs to be explored further, as many harvested populations are not monitored (i.e., freshwater crayfish). With aquatic species populations declining at a rate of 4% a year, responses to the pressures of animal product consumption are imperative. This review found that decreasing or eliminating consumption of animal products was the most impactful strategy for alleviating pressures placed on aquatic biodiversity when compared to supply-side interventions, such as increasing efficiency. More specifically, global adoption of plant-based diets could reduce agricultural land-use by 76%, which is equivalent to an area more than three times the size of the United States, including Alaska,

Hawaii, and Puerto Rico, as well as result in 358-743 GtCO2 removal via carbon sequestrations through restoration, which is roughly equivalent to the past 9 to 16 years of fossil fuel emissions.

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