Oceans in Peril: Grand Challenges in Applied Water Quality Research for the 21st Century

Microbial pollution

Microbial pollution of coastal waters occurs when enteric microbes enter waters used for recreation or aquaculture. Enteric microbes include pathogenic viruses, bacteria, and protozoa that are transmitted via the fecal-oral route, and fecal indicator organisms like Escherichia coli and enterococci typically used to assess microbial water quality. Sources of these organisms to coastal waters include land-based runoff, spilled sewage, animal (including human) feces, and treated wastewater. Microbial pollutants are routinely monitored in coastal waters of developed countries, but are not monitored in most developing countries. Shuval (2003) estimated that globally there are 120 million cases of gastrointestinal illness and 50 million cases of severe respiratory illness per year caused by swimming in microbe-polluted coastal waters. In Southern California alone it is estimated that 1.5 million cases of excess gastrointestinal illness at a cost of $50 million occur each year (Given et al., 2006). Illnesses caused by consuming filter-feeding shellfish that can accumulate pathogens are not included in the estimates. Microbial pollution in the marine environment has been detected in areas as remote as Antarctica (Delille and Delille, 2000).

In the United States and other developed regions, national regulations aim to reduce incidence of thalassogenic disease by dictating allowable levels of microbial pollution at recreational beaches (Bartram and Rees, 2000). The BEACH Act (2000) mandates that U.S. beaches be monitored for microbial pollution to protect swimmers from unacceptable levels of gastroenteritis risk. Routine monitoring resulted in 20120 beach advisories and closures in 2012 as a result of elevated levels of microbial pollutants. This number is up from 6200 in 1999 (Dorfman and Haren, 2013). Of the advisories and closures in 2012, 63% were due to unknown sources of contamination (Dorfman and Haren, 2013). At the same time, nearly one-third of the shellfish harvesting waters in the United States are classified as being “fecal impaired” by the National Shellfish Registry (NOAA, 1998). There is a clear need to identify the sources of contamination and remediate microbial pollution in coastal waters.

The USEPA recently refined its beach water quality criteria (USEPA, 2012), allowing beach managers to develop their own “site-specific” water quality criteria. Beach managers can develop site-specific criteria by showing that their beaches would be better managed using different fecal indicator organisms (USEPA, 2014) from those recommended by USEPA (E. coli and enterococci) or showing that the source of microbial pollution at the beach is different from the source assumed by USEPA (disinfected wastewater treatment plant discharge mixed with some raw sewage) in developing their water quality criteria (Soller et al., 2014). At the same time, USEPA will allow managers to use models to predict when a beach should be under a microbial pollution advisory (USEPA, 2012). There are a number of applied research needs related to microbial source tracking (MST) and modeling that can inform the implementation of these new provisions allowed by USEPA.

MST is the tracking of microbial pollution in the environment to its sources (Santo Domingo et al., 2007). Fecal indicator organisms can come from a variety of fecal (Layton et al., 2010) and nonfecal sources (Yamahara et al., 2007; Russell et al., 2013). Understanding the source of contamination can allow remedial actions to be taken, total maximum daily loads (TMDLs) to be established, health risk to be inferred, and site-specific water quality criteria to be developed (USEPA, 2012). A number of diverse MST targets have been developed. Most of these are genetic markers from bacteria that are unique to the intestinal tract of specific animals. For example, there are MST markers for humans (Shanks et al., 2009, 2010; Layton et al., 2013), dogs (Schriewer et al., 2013), birds (Lu et al., 2008; Lee et al., 2013; Sinigalliano et al., 2013), and ruminants (Shanks et al., 2008). The specificity and sensitivity of these markers appear to vary by geographic region in some cases, but they can be very high (Boehm et al., 2013).

The challenge is interpreting the concentrations of MST markers in the environment for microbial source allocation and health risk assessment. The number of MST markers relative to fecal indicator organisms varies by fecal source (Ervin et al., 2013), so the relative concentration of, for example, human and gull-associated markers cannot be used to infer which source contributes more fecal indicator organisms to a waterbody. Additionally, the decay of the MST markers relative to each other and fecal indicator organisms varies, adding additional complexity to interpreting their concentrations in water samples (Wang et al., 2013). Although microbial pollution from human feces is likely to present the greatest health risk to swimmers (Soller et al., 2014), there has been limited research to relate MST marker concentration to health risk (Harwood et al., 2014; Boehm et al., 2015). Levels of human-associated markers close to assay detection limits have been observed at urban beaches or beaches with known sewage inputs (Santoro and Boehm, 2007; Boehm et al., 2009b; Russell et al., 2013). The presence of low levels of markers has been useful for source attribution, but it is unclear whether the low level of markers might indicate a significant health risk. Research is needed to understand how concentrations of the MST targets relate to a health risk. This could be achieved with quantitative microbial risk assessments (Boehm et al., 2009a, 2015; Viau et al., 2011; Soller et al., 2014) or epidemiology studies (Colford et al., 2007, 2012; Boehm and Soller, 2011).

Research on fate and transport of microbial pollutants including pathogens, MST markers, and indicator organisms is needed. Understanding various fate and transport processes can inform the creation of models both for beach management (these tend to be statistical models (Francy, 2009; Thoe et al., 2015)) and for source allocation and TMDL development (these tend to be deterministic models) (Liu et al., 2006; Boehm et al., 2009b; Nevers and Boehm, 2010). Sunlight appears to be one of the most important factors affecting microbial pollutant concentrations in clear water (Boehm et al., 2009b), but most research has been done with indicator organisms, and very little on actual pathogens and MST markers. The roles of grazing (Boehm et al., 2005) and particle-microbe interactions in attenuating concentrations in the environment are not particularly well understood (Nevers and Boehm, 2010). A major challenge in modeling microbial pollutants in environmental waters is dealing with the inherent uncertainty (Gronewold and Wolpert, 2008), and short-period (Boehm, 2007) and small-spatial variability (Boehm et al., 2002) that has been documented in their concentrations. An open question: how precise can models really be given these measurement uncertainties?

Finally, low impact development (LID) is gaining popularity as a means for reducing the impact of microbial pollution from stormwater and urban runoff discharges on coastal waters. According to the National Resources Defense Council (Dorfman and Haren, 2013), untreated stormwater is the most frequently cited cause of beach advisories and closures in the United States. LID for pollutant reduction typically consists of passive structures, as recommended through best management practices (BMPs), such as wetlands or biofilters that treat runoff via natural processes (sunlight or passage through soil or other permeable media). Leisenring et al. (2012) reviewed the performance of BMPs around the world and found that most of them were not capable of removing fecal indicator bacteria efficiently. However, there has been very limited or no work to examine removal of indicator organisms, pathogens, or MST markers from runoff (Li et al., 2012; Mohanty et al., 2014) using BMPs. Innovations in natural and engineered treatment systems for treating stormwater or urban runoff before it discharges to the coast are needed to control inputs of microbial pollutants (Fletcher et al., 2008; Jiang et al., 2015).

Nutrient pollution

Excess nutrient inputs to coastal waters can lead to eutrophication that can dramatically affect marine food webs and ecosystem health. Sources of nutrients to coastal waters include wastewater discharges, agricultural, urban and stormwater runoff, atmospheric deposition, and groundwater. Primary and secondary changes in coastal waters characteristic of eutrophication include the following: increases in algal blooms, shifts in the dominant phytoplankton including the emergence of harmful algae, reduction in the abundance of vascular plants, benthic organisms, and pelagic fish, reduction in dissolved oxygen concentrations, and decreases in pH (Cloern, 2001; Sunda and Cai, 2012). The effects of excess nutrient inputs can be observed throughout the world, on all continents except Antarctica (Cloern, 2001; Sunda and Cai, 2012). A national assessment of estuarine waters in the United States indicates 65% of U.S. estuaries are moderately to severely eutrophic (Bricker et al., 1999). The hypoxic “dead zone” in the Gulf of Mexico is one of the most striking examples of the adverse effects of eutrophication in marine waters (Rabalais et al., 2002).

Primary productivity in marine waters is typically nitrogen limited (Howarth, 1988; Cloern, 2001). The marine nitrogen cycle is complex and involves numerous microorganisms with diverse metabolisms (Gruber, 2008). Denitrification and anaerobic ammonium oxidation (anammox) are two microbial mediated processes capable of removing dissolved nitrogen from the water column (Devol, 2008) so these processes are key in controlling total nitrogen concentrations. Recent work has identified that phosphorous limitation can also occur in marine waters (Krom et al., 1991; Rabalais et al., 2002), and in some systems nutrient co-limitation occurs (Arrigo, 2005; Saito et al., 2008). New insights regarding the marine nitrogen cycle are occurring all the time (Santoro et al., 2011). Continued fundamental research to understand the organisms active in the nitrogen cycle, the conditions under which they are active, and how to accelerate nitrogen removal in systems with excess nitrogen is needed.

Unlike microbial pollutants, there are no nationwide criteria for nutrients in marine waters; however, there exists guidance on how site-specific nutrient criteria should be created (USEPA, 2001). National criteria for nutrients are inappropriate as marine waters have nonlinear responses to nutrient inputs and concentrations (Cloern, 2001). For example, Chesapeake Bay and San Francisco Bay have historically had the same mean annual nitrogen and phosphorous concentrations, yet Chesapeake Bay is eutrophic and San Francisco Bay is not (Cloern, 2001). This is attributed to the abundance of benthic filter feeders in San Francisco Bay that are able to clear the water column of phytoplankton (Cloern, 1996). In general, the response of marine waters to nutrient addition can be modulated by tide range, stratification, residence time, density of shellfish, and riverflow (Gilbert et al., 2010). Efforts are underway in many states to develop numeric nutrient criteria for their marine waters that take into account these modulators (USEPA, 2015). Both fundamental and applied research is needed to inform the setting of numerical nutrient limits including a better understanding of how coastal waters respond to nutrient inputs. The current understanding is very site specific and the unique attributes of each estuary or coastal water appear to control its response to nutrient inputs. Widely applicable models for understanding nutrient effects in marine waters would be extremely useful.

It has been argued that eutrophic marine systems are more likely to experience blooms of harmful algae (Cloern, 2001). This is, in part, because a higher N:P or N:Si ratio in marine waters can favor dinoflagellates (Heisler et al., 2008), and dinoflagellates represent a large majority of harmful algae. Harmful algae include dinoflagellates, diatoms, and cyanobacteria that produce toxins including saxitoxins, okadaic acid, venerupin, brevetoxin, domoic acid, ciguatoxin, pfisteria toxin, and β-N-methylamino-L-alanine (BMAA) (Boehm and Bischel, 2011). Some coastal waters and/or fish and shellfish growing in the waters are routinely monitored for the presence of these toxins (i.e., Puget Sound, WA) and safety thresholds are enforced in some states (i.e., Washington). However, there are no nationwide water quality criteria for harmful algae or their toxins at the present time. Monitoring for harmful algae and their toxins is not routine in U.S. coastal waters. Research efforts to simplify or automate these measurements could increase monitoring efforts. For example, Monterey Bay Aquarium Research Institute (MBARI) has developed an environmental sample processor (ESP) that can be deployed in the ocean and monitors water for harmful algae in situ (Scholin et al., 2009; Yamahara et al., 2015).

Eutrophication and resulting hypoxia, particularly in a stratified water column, can lead to a decrease in pH due to proton production during respiration and reduced buffering capacity (Sunda and Cai, 2012). There are concerns that the decrease in pH can locally exacerbate the effect of OA caused by rising global atmospheric CO₂ concentrations (Doney et al., 2009). Coastal resource managers are increasingly interested in understanding whether reductions in nutrient discharges, from wastewater, for example, to coastal waters can alleviate problems of decreasing pH (Washington State Blue Ribbon Panel, 2012; Boehm et al., 2015). Sunda and Cai (2012) showed that in the Gulf of Mexico and the South China Sea, discharge of nutrient-laden runoff from the Mississippi and Yellow rivers, respectively, contributes to local acidification by enhancing primary productivity in stratified waters. The importance of this process in other coastal regions needs to be evaluated to determine whether local management actions by water quality managers at the scale of counties and cities can reduce the potential for OA in local waters (Boehm et al., 2015). Both monitoring and modeling efforts to assess the effects of nutrient and organic discharges, from anthropogenic waste streams and natural discharges like rivers and creeks, on local marine carbonate chemistry is needed. Advances in monitoring approaches are particularly important as glass electrodes typically used to measure pH in coastal seawater by dischargers for their monitoring programs are not precise enough to discern small changes in coastal pH (Boehm et al., 2015). Newer sensors are becoming available to measure other relevant carbonate system parameters like pCO₂ and alkalinity (Byrne, 2014) and additional efforts to develop such sensors are needed.

Engineered and natural systems can be designed to remove nutrients from waste flows. Biological denitrification, anammox, and chemical or biological phosphorous unit processes can remove nutrients from wastewater flows (Tchobanoglous et al., 2003). Upgrading wastewater treatment facilities to remove nutrients using these or other advanced treatments may be necessary to reduce nutrient fluxes to sensitive habitats, particularly once nutrient criteria are developed. Natural treatment systems such as biofilters (Bratieres et al., 2008), wetlands (Jasper et al., 2014), and grass swales can remove nutrients from urban and stormwater runoff, although their performance can be unreliable (Fletcher et al., 2008; Leisenring et al., 2012; Li and Davis, 2014). Work that informs the design of engineered and natural treatment systems that effectively and predictably remove nutrients from runoff in particular is needed to reduce nutrient inputs to coastal waters. For example, recent studies illustrated that addition of newspapers or woodchips to biofilters can enhance denitrification (Goh et al., 2015; Peterson et al., 2015).

Chemical contaminants

A variety of chemical contaminants have been extensively studied in the marine environment, including organochlorine compounds (i.e., polychlorinated biphyenls [PCBs] and pesticides), polyaromatic hydrocarbons (PAHs), heavy metals, and radionuclides. These contaminants are usually strongly associated with particulate matter and bioaccumulate in organisms, biomagnify in the foodweb, and negatively affect ecosystem health. Adverse health effects from exposure to these contaminants range from disruption of the endocrine system to cancer in a variety of organisms, including humans (Addison, 1996; Longnecker et al., 1997; Gilbert et al., 2002; Hoeve and Jacobson, 2012). While some of these chemicals can naturally occur in the environment, elevated concentrations are typically found in coastal areas due to anthropogenic inputs from agricultural, urban, and industrial activities. The contaminants enter the environment from both terrestrial runoff and atmospheric deposition and undergo environmental cycling (between sediment, air, and land), with concentrations found even in remote regions of the world such as the Arctic (Muir et al., 1992; Ferm, 1996).

Regulation of the production, use, and discharge of organochlorine compounds, polyaromatic hydrocarbons, heavy metals, and radionuclides has reduced their input into the environment, but their persistence has continued to make it challenging to effectively remediate them. For example, approximately 6.1 million metric tons of petroleum products containing PAHs are annually released into the ocean primarily due to anthropogenic sources, and this number excludes single catastrophic events (Haynes and Johnson, 2000). While dispersants are used to cleanup petroleum products in marine systems to reduce immediate impacts to wildlife, the dispersants themselves alter the behavior of the hydrocarbons, resulting in increased water solubility and bioavailability, which can cause long-term delayed toxicological impacts to marine organisms from exposure (Wolfe et al., 1998). Similar to PAHs, organochlorine, heavy metal, and radionuclide sorption and solubility behavior can change in different environmental conditions altering environmental fate and bioavailability over time (Waldichuk, 1985; Rainbow, 1995; Haynes and Johnson, 2000; Buesseler et al., 2011; Yoshida and Kanda, 2012). Due to the behavior and persistence of these compounds, solutions to reduce their inputs into the environment are needed, as are techniques for cost-effective large-scale remediation after inadvertent inputs. Conventional engineering approaches that employ a single technology may be cost-intensive and less effective than combining multiple remediation strategies. For example, using electrobioremediaton, which combines electrokinetic remediation and bioremediation, accelerates degradation and improves removal efficiency of hydrophobic organic contaminants in comparison to using a single technology (Li et al., 2010; Megharaj et al., 2011). Further work to combine physical, chemical, and biological remediation strategies has significant potential to result in novel treatment approaches.

While the sources, fate, and health effects of the above compounds are relatively well known, a class of contaminants requiring further characterization is contaminants of emerging concern (CECs). CECs include a large number of compounds with a variety of uses and physicochemical properties. Common classes of CECs are pharmaceuticals, personal care products, pesticides, herbicides, perfluorinated compounds (PFCs), and flame-retardants. CECs found in coastal waters are from both point and nonpoint sources including wastewater effluent, industrial waste, stormwater, and runoff. Due to the large number of chemicals classified as CECs with different properties and the lack of established measurement methods for many of these compounds, the source, fate, transport, and effects of CECs in the marine environment are not well understood. Preliminary studies show that many CECs can be detected in coastal marine waters that are impacted by wastewater discharge, in sediment, and in fish (Maruya et al., 2012; Vidal-Dorsch et al., 2012). Toxicity studies have also shown that certain pharmaceuticals can lead to endocrine and physiological disruption in marine organisms, but the toxicities at environmentally relevant concentrations of many individual CECs, or mixtures of them that can have synergistic effects, have not been measured (Bay and Vidal‐Dorsch, 2012; Scott et al., 2012). A 2-year pilot study, focused on quantifying CECs in mussel tissue in coastal regions in California, prioritized further monitoring of four types of CECs due to the high frequency of detection and concentration of these contaminants (Maruya et al., 2014). The CECs recommended for future monitoring efforts are PFCs, flame retardants, the additive 4-nonylphenol and its derivatives, and the pharmaceutical lomefloxacin (Maruya et al., 2014). While the California-based CEC study provides initial guidance on monitoring priorities, additional studies are needed on a larger spatial and temporal scale to confirm its findings and determine priority CECs based on location and associated land use. With the exception of polybrominated flame retardants and PFCs, very limited information exists on CECs in the marine environment beyond the California coast. Table 2 summarizes findings from relevant studies on personal care products and pharmaceuticals (PPCPs) in the marine environment. Based on current review of available published literature, analysis of PPCPs in the marine environment is limited to coastal environments and does not include samples from remote regions.

For CECs, an integrated approach of preventing and characterizing chemical inputs into the environment is necessary. Studies have shown that marine sampling locations impacted by stormwater discharge had higher detection frequency and concentrations of target CECs than those impacted primarily by wastewater discharge from publicly owned treatment works (POTWs) (Maruya et al., 2012, 2014; Alvarez et al., 2014). Thus, managing stormwater and urban runoff using LID, such as bioretention ponds (Hsieh and Davis, 2005) as well as improving conventional wastewater treatment through advanced treatment, such as ozonation and wetlands (Oulton et al., 2010), can reduce the inputs of CECs to coastal waters. The types of CECs discharged from POTW effluent and runoff can vary based on regional and industrial activities, which will impact the effectiveness of advanced treatment or LID. Monitoring of CECs is needed to develop a better understanding of the sources and behavior of these contaminants undergoing treatment.

Developing cost-effective analytical tools and methods to measure mixtures of CECs in different matrices is needed to better characterize environmental concentrations. Better detection methods to measure trace levels of these contaminants will allow study of fate and transport and of ecotoxicological and human health effects at environmentally relevant concentrations. Analytical tools that expedite detection of these contaminants in a variety of organisms at environmentally relevant concentrations can facilitate toxicity studies to increase understanding of the synergistic and chronic health effects of these contaminants and develop appropriate sublethal toxicological parameters (Fent et al., 2006). Advances in analytical equipment such as ultra-high-performance liquid chromatography has resulted in significant progress in detecting CECs, but shortcomings include high cost and lack of validation procedures for different analytical methods and environmental samples (Wille et al., 2012). Biosensors, which can provide real-time data on the concentrations and biological effects of CECs, are a potential alternative to conventional techniques of liquid chromatography and bioassays, but most CEC biosensors are still in developmental stages in the laboratory and have not been deployed in the field and are not commercially available (Rodriguez-Mozaz et al., 2005). Sampling studies have excluded certain CECs, such as phalates and alkylphenols, due to lack of confidence in performance and reliability of current analytical methods (Bay and Vidal‐Dorsch, 2012; Maruya et al., 2012; Scott et al., 2012). Due to the lack of studies quantifying CECs in the marine environment it is not yet possible to determine the environmental risk associated with most of these contaminants (Scott et al., 2012). Combining advanced analytical tools with improved biomarker monitoring can result in cost- and time-effective characterization of these CECs in water, sediment, and biota, which in turn can lead to risk assessments that better inform policy.

Marine debris (plastics)

Marine debris, defined as anthropogenic material such as plastic that enters and persists in the marine environment, represents a growing problem (Cole et al., 2011; Engler, 2012; Wright et al., 2013). Marine plastics are just one type of internationally recognized highly persistent pollutant that may have a significant effect on the ecological integrity of the ocean and ultimately affect human health (Gregory, 2009; Lozano and Mouat, 2009). Marine plastics of varying sizes are globally yet heterogeneously distributed, and it is estimated that up to 10% of the 265 million tons of plastics produced annually eventually enter the marine environment (Cole et al., 2011; Gouin et al., 2011; Bakir et al., 2014). Plastics enter oceans via a variety of sources, with 80% of marine plastics generated from land-based sources, such as beach litter and surface runoff (Andrady, 2011; Cole et al., 2011). Plastic debris ranges in size from discarded fishing nets to tiny pieces of plastics from air-blasting or scrubbing soaps. While plastics are persistent in the environment, they can break down over time into smaller fragments due to chemical, physical, and biological weathering (Cole et al., 2011; Koelmans et al., 2013). Once in the ocean, plastic debris can be detrimental to marine life in various ways. Large plastic debris such as fishing lines and nets pose a threat to marine life entanglement, while smaller plastic debris often referred to as microplastics, can be ingested by marine organisms.

Microplastics, defined as plastics having a particle size less than 5 mm, have recently been identified as CECs, and are considered a leading global environmental issue (Browne et al., 2007; Sutherland et al., 2010; Andrady, 2011; Cole et al., 2011; Wagner et al., 2014). Microplastic abundance has increased by two orders of magnitude over the past 40 years in the North Pacific, and the abundance of microplastics in all marine environments is projected to continue to increase (Wright et al., 2013). Microplastics can cycle through the food web and enter the human food chain (Wright et al., 2013). They are transported long distances, reaching remote regions of the globe, are ingested by aquatic organisms, and then transferred between trophic levels (Andrady, 2011; Cole et al., 2011; Wright et al., 2013). Microplastic ingestion and retention occurs in a wide range of marine aquatic organisms at varying trophic levels from zooplankton to fish, and trophic transfer has also been shown to occur among different organisms (Browne et al., 2007; Graham and Thompson, 2009; Murray and Cowie, 2011; Cole et al., 2013; Farrell and Nelson, 2013; Chua et al., 2014; Setälä et al., 2014; Van Cauwenberghe and Janssen, 2014). Accumulated microplastics in bivalves (Van Cauwenberghe and Janssen, 2014) and decapod crustaceans (Murray and Cowie, 2011) used for human consumption have been documented, indicating the potential for human intake of microplastics through dietary exposure.

In addition to the microplastic itself acting as a potentially harmful contaminant, microplastics act as a source and sink of both organic chemicals and trace metal contaminants (Andrady, 2011; Engler, 2012). Hydrophobic organic contaminants and trace metals sorb onto microplastics in the environment and concentrate to higher levels than the ambient seawater. Sorption of these contaminants can increase over time as the plastic ages. Marine plastics that undergo aging exhibit changes in porosity, which facilitates subsequent organic matter fouling, thereby altering the pellets' surface area and polarity, potentially leading to increased sorption (Rios et al., 2007; Ashton et al., 2010; Mato et al., 2014; Rochman et al., 2014). Studies have also shown these contaminants can then be transferred to marine species once ingested, but the extent to which the contaminants affect the ecosystem has yet to be fully understood, with conflicting assessments of environmental exposure (Teuten et al., 2007; Bakir et al., 2014).

A better understanding of the fate, transport, and impacts of marine plastics, and specifically microplastics, is necessary when considering the most effective approaches to take to remediate and prevent these pollutants. Although reported values from decades of sea surface samples provide an initial estimate of the abundance and distribution trends of microplastics in the open ocean and gyres, more extensive research on microplastic abundance and distribution in coastal systems is needed because initial studies indicate that coastal areas near industrial and urban centers can be hotspots (Browne et al., 2011; Norén and Naustvoll, 2011; Wright et al., 2013). Unfortunately, given the current available technologies, global removal of plastics currently in the environment is an insurmountable undertaking. Various ocean “garbage patches” in the subtropical gyres, such as the Great Pacific Garbage patch, have concentrations of marine plastics orders of magnitude higher than the rest of the open ocean (Law et al., 2010). Attempting to clean up only plastic in the Great Pacific Garbage Patch would not only require use of numerous ships and large amounts of fuel, but also would need to account for spatial variability and heterogeneity of the Garbage Patch, and minimize inadvertent disturbance to marine life and removal of pelagic organisms (Moore et al., 2001; Goldstein et al., 2013). While new techniques to sample, detect, and analyze microplastics in the marine environment have been developed (Hidalgo-Ruz et al., 2012; Nuelle et al., 2014), published research on remediation techniques for plastics is lacking. To combat the projected increase in marine plastic pollution, innovative technologies to prevent plastics from entering the marine environment and to degrade plastics already in the environment need to be developed. Calculations completed by Jambeck et al. (2015) show that mismanaged waste is a major source of land-based input of microplastics in the marine environment and that top contributors are countries with fast economic growth with lagging waste management infrastructure. Potential approaches include development of technologies that implement BMPs that employ physical barriers to prevent inputs into the marine environment and the use of microbial populations capable of degrading microplastics (Shah et al., 2008). In addition, implementation of biodegradable plastics on a large scale to replace current plastics used in industry is needed. The use of biodegradable plastics such as polyhydroxyalkanoates (PHA) is currently limited by high production cost using the traditional processing of PHA from cultivated feedstocks from corn and sugar cane (Rostkowski et al., 2012). Research examining the production of PHAs using the waste biogas methane and methanotrophic bacteria as an alternative to the traditional approach of PHA production, shows promise of reduced economic and environmental costs (Pieja et al., 2011; Rostkowski et al., 2012). Continuation of research in innovative production and application of biodegradable plastic can lead to cost-effective replacement of traditional plastics.