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<!doctype html> <html data-n-head-ssr lang="en" data-n-head="%7B%22lang%22:%7B%22ssr%22:%22en%22%7D%7D"> <head > <link data-n-head="ssr" rel="icon" type="image/png" sizes="16x16" href="https://brand.frontiersin.org/m/ed3f9ce840a03d7/favicon_16-tenantFavicon-Frontiers.png"> <link data-n-head="ssr" rel="icon" type="image/png" sizes="32x32" href="https://brand.frontiersin.org/m/ed3f9ce840a03d7/favicon_32-tenantFavicon-Frontiers.png"> <link data-n-head="ssr" rel="apple-touch-icon" type="image/png" sizes="180x180" href="https://brand.frontiersin.org/m/ed3f9ce840a03d7/favicon_180-tenantFavicon-Frontiers.png"> <title>Frontiers | The Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary</title><meta data-n-head="ssr" charset="utf-8"><meta data-n-head="ssr" name="viewport" content="width=device-width, initial-scale=1"><meta data-n-head="ssr" data-hid="charset" charset="utf-8"><meta data-n-head="ssr" data-hid="mobile-web-app-capable" name="mobile-web-app-capable" content="yes"><meta data-n-head="ssr" data-hid="apple-mobile-web-app-title" name="apple-mobile-web-app-title" content="Frontiers | Articles"><meta data-n-head="ssr" data-hid="theme-color" name="theme-color" content="#0C4DED"><meta data-n-head="ssr" data-hid="description" property="description" name="description" content="Engineered aeration is one solution for increasing oxygen concentrations in highly eutrophic estuaries that undergo seasonal hypoxia. Although there are vari..."><meta data-n-head="ssr" data-hid="og:title" property="og:title" name="title" content="Frontiers | The Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary"><meta data-n-head="ssr" data-hid="og:description" property="og:description" name="description" content="Engineered aeration is one solution for increasing oxygen concentrations in highly eutrophic estuaries that undergo seasonal hypoxia. Although there are vari..."><meta data-n-head="ssr" data-hid="keywords" name="keywords" content="Methane,Aeration,Eutrophication,estuary,OsmoSampler"><meta data-n-head="ssr" data-hid="og:site_name" property="og:site_name" name="site_name" content="Frontiers"><meta data-n-head="ssr" data-hid="og:image" property="og:image" name="image" content="https://images-provider.frontiersin.org/api/ipx/w=1200&f=png/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g001.jpg"><meta data-n-head="ssr" data-hid="og:type" property="og:type" name="type" content="article"><meta data-n-head="ssr" data-hid="og:url" property="og:url" name="url" content="https://www.frontiersin.org/journals/environmental-science/articles/10.3389/fenvs.2022.866152/full"><meta data-n-head="ssr" data-hid="twitter:card" name="twitter:card" content="summary_large_image"><meta data-n-head="ssr" data-hid="citation_volume" name="citation_volume" content="10"><meta data-n-head="ssr" data-hid="citation_journal_title" name="citation_journal_title" content="Frontiers in Environmental Science"><meta data-n-head="ssr" data-hid="citation_publisher" name="citation_publisher" content="Frontiers"><meta data-n-head="ssr" data-hid="citation_journal_abbrev" name="citation_journal_abbrev" content="Front. Environ. Sci."><meta data-n-head="ssr" data-hid="citation_issn" name="citation_issn" content="2296-665X"><meta data-n-head="ssr" data-hid="citation_doi" name="citation_doi" content="10.3389/fenvs.2022.866152"><meta data-n-head="ssr" data-hid="citation_firstpage" name="citation_firstpage" content="866152"><meta data-n-head="ssr" data-hid="citation_language" name="citation_language" content="English"><meta data-n-head="ssr" data-hid="citation_title" name="citation_title" content="The Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary"><meta data-n-head="ssr" data-hid="citation_keywords" name="citation_keywords" content="Methane; Aeration; Eutrophication; estuary; OsmoSampler"><meta data-n-head="ssr" data-hid="citation_abstract" name="citation_abstract" content="<p>Engineered aeration is one solution for increasing oxygen concentrations in highly eutrophic estuaries that undergo seasonal hypoxia. Although there are various designs for engineered aeration, all approaches involve either destratification of the water column or direct injection of oxygen or air through fine bubble diffusion. To date, the effect of either approach on estuarine methane dynamics remains unknown. Here we tested the hypotheses that 1) bubble aeration will strip the water of methane and enhance the air-water methane flux to the atmosphere and 2) the addition of oxygen to the water column will enhance aerobic methane oxidation in the water column and potentially offset the air-water methane flux. These hypotheses were tested in Rock Creek, Maryland, a shallow-water sub-estuary to the Chesapeake Bay, using controlled, ecosystem-scale deoxygenation experiments where the water column and sediments were sampled in mid-summer, when aerators were ON, and then 1, 3, 7, and 13 days after the aerators were turned OFF. Experiments were performed under two system designs, large bubble and fine bubble approaches, using the same observational approach that combined discrete water sampling, long term water samplers (OsmoSamplers) and sediment porewater profiles. Regardless of aeration status, methane concentrations reached as high as 1,500 nmol L<sup>−1</sup> in the water column during the experiments and remained near 1,000 nmol L<sup>−1</sup> through the summer and into the fall. Since these concentrations are above atmospheric equilibrium of 3 nmol L<sup>−1</sup>, these data establish the sub-estuary as a source of methane to the atmosphere, with a maximum atmospheric flux as high as 1,500 µmol m<sup>−2</sup> d<sup>−1</sup>, which is comparable to fluxes estimated for other estuaries. Air-water methane fluxes were higher when the aerators were ON, over short time frames, supporting the hypothesis that aeration enhanced the atmospheric methane flux. The fine-bubble approach showed lower air-water methane fluxes compared to the larger bubble, destratification system. We found that the primary source of the methane was the sediments, however, <italic>in situ</italic> methane production or an upstream methane source could not be ruled out. Overall, our measurements of methane concentrations were consistently high in all times and locations, supporting consistent methane flux to the atmosphere.</p>"><meta data-n-head="ssr" data-hid="citation_pdf_url" name="citation_pdf_url" content="https://www.frontiersin.org/journals/environmental-science/articles/10.3389/fenvs.2022.866152/pdf"><meta data-n-head="ssr" data-hid="citation_online_date" name="citation_online_date" content="2022/06/13"><meta data-n-head="ssr" data-hid="citation_publication_date" name="citation_publication_date" content="2022/08/11"><meta data-n-head="ssr" data-hid="citation_author_0" name="citation_author" content="Lapham, Laura L."><meta data-n-head="ssr" data-hid="citation_author_institution_0" name="citation_author_institution" content="Chesapeake Biological Laboratory, United States"><meta data-n-head="ssr" data-hid="citation_author_1" name="citation_author" content="Hobbs, Edward A."><meta data-n-head="ssr" data-hid="citation_author_institution_1" name="citation_author_institution" content="Chesapeake Biological Laboratory, United States"><meta data-n-head="ssr" data-hid="citation_author_2" name="citation_author" content="Testa, Jeremy M."><meta data-n-head="ssr" data-hid="citation_author_institution_2" name="citation_author_institution" content="Chesapeake Biological Laboratory, United States"><meta data-n-head="ssr" data-hid="citation_author_3" name="citation_author" content="Heyes, Andrew"><meta data-n-head="ssr" data-hid="citation_author_institution_3" name="citation_author_institution" content="Chesapeake Biological Laboratory, United States"><meta data-n-head="ssr" data-hid="citation_author_4" name="citation_author" content="Forsyth, Melinda K."><meta data-n-head="ssr" data-hid="citation_author_institution_4" name="citation_author_institution" content="Chesapeake Biological Laboratory, United States"><meta data-n-head="ssr" data-hid="citation_author_5" name="citation_author" content="Hodgkins, Casey"><meta data-n-head="ssr" data-hid="citation_author_institution_5" name="citation_author_institution" content="Chesapeake Biological Laboratory, United States"><meta data-n-head="ssr" data-hid="citation_author_6" name="citation_author" content="Szewczyk, Curtis"><meta data-n-head="ssr" data-hid="citation_author_institution_6" name="citation_author_institution" content="Chesapeake Biological Laboratory, United States"><meta data-n-head="ssr" data-hid="citation_author_7" name="citation_author" content="Harris, Lora A."><meta data-n-head="ssr" data-hid="citation_author_institution_7" name="citation_author_institution" content="Chesapeake Biological Laboratory, United States"><meta data-n-head="ssr" data-hid="dc.identifier" name="dc.identifier" content="doi:10.3389/fenvs.2022.866152"><link data-n-head="ssr" rel="manifest" href="/article-pages/_nuxt/manifest.c499fc0a.json" data-hid="manifest"><link data-n-head="ssr" rel="canonical" href="https://www.frontiersin.org/journals/environmental-science/articles/10.3389/fenvs.2022.866152/full"><script data-n-head="ssr" data-hid="newrelic-browser-script" type="text/javascript">window.NREUM||(NREUM={});NREUM.info = {"agent":"","beacon":"bam.nr-data.net","errorBeacon":"bam.nr-data.net","licenseKey":"598a124f17","applicationID":"588603994","agentToken":null,"applicationTime":2.620559,"transactionName":"MQcDMkECCkNSW0YMWghNIgldDQFTRxd1IGFJTQ==","queueTime":0,"ttGuid":"13a033963aae4a1c"}; (window.NREUM||(NREUM={})).init={privacy:{cookies_enabled:true},ajax:{deny_list:["bam.nr-data.net"]},distributed_tracing:{enabled:true}};(window.NREUM||(NREUM={})).loader_config={agentID:"594400880",accountID:"230385",trustKey:"230385",xpid:"VgUHUl5WGwYIXFdSBAgOUg==",licenseKey:"598a124f17",applicationID:"588603994"};;/*! 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McGinnis" class="Avatar__img is-inside-mask"></figure> <div class="ArticleDetailsEditors__ediorInfo__info"><div class="ArticleDetailsEditors__ediorInfo__name"> Daniel F. McGinnis </div> <div class="ArticleDetailsEditors__ediorInfo__affiliation"> University of Geneva, Switzerland </div></div></a></div></div> <div class="ArticleDetailsEditors"><div class="ArticleDetailsEditors__editors"><div class="ArticleDetailsEditors__title">Reviewed by</div> <a href="https://loop.frontiersin.org/people/218446/overview" data-event="editorInfo-a-weiDongZhai" class="ArticleDetailsEditors__ediorInfo"><figure class="Avatar Avatar--size-32"><img src="https://loop.frontiersin.org/images/profile/218446/32" alt="Wei-dong Zhai" class="Avatar__img is-inside-mask"></figure> <div class="ArticleDetailsEditors__ediorInfo__info"><div class="ArticleDetailsEditors__ediorInfo__name"> Wei-dong Zhai </div> <div class="ArticleDetailsEditors__ediorInfo__affiliation"> Southern Marine Science and Engineering Guangdong Laboratory (Zhuhai), China </div></div></a><a href="https://loop.frontiersin.org/people/1668703/overview" data-event="editorInfo-a-peterTango" class="ArticleDetailsEditors__ediorInfo"><figure class="Avatar Avatar--size-32"><img src="https://loop.frontiersin.org/images/profile/1668703/32" alt="Peter Tango" class="Avatar__img is-inside-mask"></figure> <div class="ArticleDetailsEditors__ediorInfo__info"><div class="ArticleDetailsEditors__ediorInfo__name"> Peter Tango </div> <div class="ArticleDetailsEditors__ediorInfo__affiliation"> United States Geological Survey (USGS), United States Department of the Interior, United States </div></div></a></div></div> <div class="ArticleDetailsGlossary ArticleDetailsGlossary--open"><button class="ArticleDetailsGlossary__header"><div class="ArticleDetailsGlossary__header__title">Table of contents</div> <div class="ArticleDetailsGlossary__header__arrow"></div></button> <div class="ArticleDetailsGlossary__content"><ul class="flyoutJournal"><li><a href="#h1">Abstract</a></li><li><a href="#h2">1 Introduction</a></li><li><a href="#h3">2 Materials and Methods</a></li><li><a href="#h4">3 Results</a></li><li><a href="#h5">4 Discussion</a></li><li><a href="#h6">5 Conclusion</a></li><li><a href="#h7">Data Availability Statement</a></li><li><a href="#h8">Author Contributions</a></li><li><a href="#h9">Funding</a></li><li><a href="#h10">Conflict of Interest</a></li><li><a href="#h11">Publisher’s Note</a></li><li><a href="#h12">Acknowledgments</a></li><li><a href="#h13">Supplementary Material</a></li><li><a href="#h14">References</a></li></ul></div></div>  <div class="ActionsDropDown"><button aria-label="Open dropdown" data-event="actionsDropDown-button-toggle" class="ActionsDropDown__button ActionsDropDown__button--typeOutline ActionsDropDown__button--iconQuote"><span class="ActionsDropDown__button__label">Export citation</span></button> <div class="ActionsDropDown__menuWrapper"> <ul class="ActionsDropDown__menu"><li><a href="/journals/environmental-science/articles/10.3389/fenvs.2022.866152/endNote" target="_blank" rel="noopener noreferrer" 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article</h2> <div class="ArticleLayoutHeader__info__journalDate"><span>Front. Environ. Sci.</span><span>, 11 August 2022</span></div> <div class="ArticleLayoutHeader__info__journalDate"> Sec. Biogeochemical Dynamics </div> <div class="ArticleLayoutHeader__info__doiVolume"><span> Volume 10 - 2022 | </span> <a href="https://doi.org/10.3389/fenvs.2022.866152" class="ArticleLayoutHeader__info__doi"> https://doi.org/10.3389/fenvs.2022.866152 </a></div> </div>  <div class="ArticleLayoutHeader__isPartOfRT"><span class="ArticleLayoutHeader__isPartOfRT__label">This article is part of the Research Topic</span> <span class="ArticleLayoutHeader__isPartOfRT__title">Physical and Biogeochemical Processes Driving Methane Sources, Sinks and Emissions in Aquatic Systems: The Past, Present and Future under Global Change</span> <span class="Link__wrapper"><a aria-label="View all 24 articles" href="https://www.frontiersin.org/research-topics/22659/physical-and-biogeochemical-processes-driving-methane-sources-sinks-and-emissions-in-aquatic-systems-the-past-present-and-future-under-global-change/articles" target="_self" data-event="customLink-link-a_viewAll24Articles" class="Link Link--linkType Link--maincolor Link--medium Link--icon Link--chevronRight Link--right"><span>View all 24 articles</span></a></span></div></div> <div class="ArticleDetails__main__content"><div class="ArticleDetails__main__content__main ArticleDetails__main__content__main--fullArticle"><div class="JournalAbstract"><div class="JournalAbstract__titleWrapper"><h1>The Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary</h1> </div> </div> <div class="JournalFullText"><div class="JournalAbstract"><a id="h1" name="h1"></a><div class="authors"><span class="author-wrapper"><a href="https://loop.frontiersin.org/people/23188" class="user-id-23188"><img class="pr5" src="https://loop.frontiersin.org/images/profile/23188/74" onerror="this.onerror=null;this.src='https://loop.frontiersin.org/cdn/images/profile/default_32.jpg';" alt="Laura L. Lapham
">Laura L. Lapham</a>*</span><span class="author-wrapper"><a href="https://loop.frontiersin.org/people/1413210" class="user-id-1413210"><img class="pr5" src="https://loop.frontiersin.org/images/profile/1413210/74" onerror="this.onerror=null;this.src='https://loop.frontiersin.org/cdn/images/profile/default_32.jpg';" alt="Edward A. Hobbs">Edward A. Hobbs</a></span><span class="author-wrapper"><a href="https://loop.frontiersin.org/people/435564" class="user-id-435564"><img class="pr5" src="https://loop.frontiersin.org/images/profile/435564/74" onerror="this.onerror=null;this.src='https://loop.frontiersin.org/cdn/images/profile/default_32.jpg';" alt="Jeremy M. Testa">Jeremy M. Testa</a></span><span class="author-wrapper"><a href="https://loop.frontiersin.org/people/1675968" class="user-id-1675968"><img class="pr5" src="https://loop.frontiersin.org/images/profile/1675968/74" onerror="this.onerror=null;this.src='https://loop.frontiersin.org/cdn/images/profile/default_32.jpg';" alt="Andrew Heyes">Andrew Heyes</a></span><span class="author-wrapper"><img class="pr5" src="https://loop.frontiersin.org/cdn/images/profile/default_32.jpg" alt="Melinda K. Forsyth" onerror="this.onerror=null;this.src='https://loop.frontiersin.org/cdn/images/profile/default_32.jpg';">Melinda K. Forsyth</span><span class="author-wrapper"><img class="pr5" src="https://loop.frontiersin.org/cdn/images/profile/default_32.jpg" alt="Casey Hodgkins" onerror="this.onerror=null;this.src='https://loop.frontiersin.org/cdn/images/profile/default_32.jpg';">Casey Hodgkins</span><span class="author-wrapper"><a href="https://loop.frontiersin.org/people/1710495" class="user-id-1710495"><img class="pr5" src="https://loop.frontiersin.org/images/profile/1710495/74" onerror="this.onerror=null;this.src='https://loop.frontiersin.org/cdn/images/profile/default_32.jpg';" alt="Curtis Szewczyk">Curtis Szewczyk</a></span><span class="author-wrapper"><a href="https://loop.frontiersin.org/people/1264726" class="user-id-1264726"><img class="pr5" src="https://loop.frontiersin.org/images/profile/1264726/74" onerror="this.onerror=null;this.src='https://loop.frontiersin.org/cdn/images/profile/default_32.jpg';" alt="Lora A. Harris">Lora A. Harris</a></span></div><ul class="notes"><li>Chesapeake Biological Laboratory, University of Maryland Center for Environmental Science, Solomons, MD, United States</li></ul><p class="mb15">Engineered aeration is one solution for increasing oxygen concentrations in highly eutrophic estuaries that undergo seasonal hypoxia. Although there are various designs for engineered aeration, all approaches involve either destratification of the water column or direct injection of oxygen or air through fine bubble diffusion. To date, the effect of either approach on estuarine methane dynamics remains unknown. Here we tested the hypotheses that 1) bubble aeration will strip the water of methane and enhance the air-water methane flux to the atmosphere and 2) the addition of oxygen to the water column will enhance aerobic methane oxidation in the water column and potentially offset the air-water methane flux. These hypotheses were tested in Rock Creek, Maryland, a shallow-water sub-estuary to the Chesapeake Bay, using controlled, ecosystem-scale deoxygenation experiments where the water column and sediments were sampled in mid-summer, when aerators were ON, and then 1, 3, 7, and 13 days after the aerators were turned OFF. Experiments were performed under two system designs, large bubble and fine bubble approaches, using the same observational approach that combined discrete water sampling, long term water samplers (OsmoSamplers) and sediment porewater profiles. Regardless of aeration status, methane concentrations reached as high as 1,500 nmol L<sup>−1</sup> in the water column during the experiments and remained near 1,000 nmol L<sup>−1</sup> through the summer and into the fall. Since these concentrations are above atmospheric equilibrium of 3 nmol L<sup>−1</sup>, these data establish the sub-estuary as a source of methane to the atmosphere, with a maximum atmospheric flux as high as 1,500 µmol m<sup>−2</sup> d<sup>−1</sup>, which is comparable to fluxes estimated for other estuaries. Air-water methane fluxes were higher when the aerators were ON, over short time frames, supporting the hypothesis that aeration enhanced the atmospheric methane flux. The fine-bubble approach showed lower air-water methane fluxes compared to the larger bubble, destratification system. We found that the primary source of the methane was the sediments, however, <em>in situ</em> methane production or an upstream methane source could not be ruled out. Overall, our measurements of methane concentrations were consistently high in all times and locations, supporting consistent methane flux to the atmosphere.</p><div class="clear"></div></div><div class="JournalFullText"><a id="h2" name="h2"></a><h2>1 Introduction</h2><p class="mb15">The eutrophication of estuaries as a result of nutrient enrichment is a global phenomenon, with consequences that include deoxygenation and hypoxia (<a href="#B19">Diaz and Rosenberg, 2008</a>). In fact, modeling and data analysis suggests that dissolved oxygen in estuaries will continue to decline into the future, primarily as a result of long-term warming (<a href="#B7">Breitburg et al., 2018</a>; <a href="#B62">Ni et al., 2019</a>; <a href="#B82">Whitney and Vlahos, 2021</a>). The primary mitigation tool has been to enforce managed reductions of land-based nutrients in the United States (<a href="#B46">Linker et al., 2013</a>) and in Europe (<a href="#B30">HELCOM, 2021</a>), yet engineered solutions are also being considered (<a href="#B14">Conley et al., 2009</a>; <a href="#B44">Lehtoranta et al., 2022</a>). Engineered aeration efforts work by either destratifying the water column or directly injecting oxygen to the water (<a href="#B28">Harris et al., 2015</a>; <a href="#B78">Stigebrandt et al., 2015</a>; <a href="#B39">Koweek et al., 2020</a>). This has also been commonly done in small lake systems (e.g., <a href="#B51">Martinez and Anderson, 2013</a>; <a href="#B32">Hounshell et al., 2021</a>) and reservoirs (<a href="#B55">McCord et al., 2016</a>). While aeration should relieve the low oxygen problem to create habitat for metazoan life, prevent the noxious release of sulfide from sediments, and enhance coupled nitrification-denitrification, an additional potential consequence is that aeration could also enhance atmospheric methane emissions in estuaries. If this is true, methane emissions from estuaries that undergo aeration could be larger than currently considered in the global budget (<a href="#B72">Saunois et al., 2020</a>). It is critical to constrain all sources of methane to the atmosphere since it is a powerful greenhouse gas (<a href="#B22">Forster et al., 2007</a>; <a href="#B20">Dlugokencky, 2020</a>). However, studies from an aerated freshwater reservoir show that the methane emissions were lower than a nearby natural reservoir (<a href="#B53">McClure et al., 2018</a>; <a href="#B54">McClure et al., 2021</a>). To date, this interplay between engineered aeration and methane fluxes in a natural estuary has not been rigorously tested.</p><p class="mb15">Estuaries are dynamic environments, generating temporally and spatially-varying habitats in which methane producing and consuming processes occur. The primary source of methane is the underlying sediments, as in most organic rich environments (<a href="#B49">Martens and Berner, 1974</a>; <a href="#B67">Reeburgh, 2007</a>). However, there is an increasing appreciation for alternative sources such as demethylation of organic phosphonates (<a href="#B37">Karl et al., 2008</a>), bacterial degradation of water column dissolved organic matter (<a href="#B68">Repeta et al., 2016</a>) and/or production by phytoplankton (<a href="#B4">Bižić et al., 2020</a>) that have not been fully explored in estuarine systems. In shallow-water, dynamic coastal and estuarine environments, methane can also be delivered with currents or tides from lateral sources (<a href="#B4">Bižić et al., 2020</a>). While methane formed in the sediments can enter the water column through ebullition (<a href="#B6">Boudreau, 2012</a>) or diffusion, nearly 85% of the methane produced within sediments is oxidized anaerobically before it reaches the sediment-water interface <em>via</em> microbially mediated reactions including sulfate reduction, nitrate reduction, and iron reduction (<a href="#B23">Froelich et al., 1979</a>; <a href="#B67">Reeburgh, 2007</a>). The remaining methane released from the sediments to the overlying water column can then be oxidized aerobically <em>via</em> methanotrophs (<a href="#B27">Hanson and Hanson, 1996</a>). Thus a conceptual model for a healthy estuary shows a small methane flux to the atmosphere (<a href="#F1">Figure 1A</a>). Alternatively, when estuarine waters are highly eutrophic, there is a breakdown in the aerobic biofilter in the water column and this results in an enhanced methane flux when the bottom waters go hypoxic and anoxic (<a href="#B24">Gelesh et al., 2016</a>). Thus, under these conditions, there is a higher methane flux to the atmosphere (<a href="#F1">Figure 1B</a>).</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 1</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g001.jpg" name="Figure1" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g001.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g001.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g001.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g001.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g001.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g001.jpg" alt="www.frontiersin.org" id="F1" loading="lazy"> </picture> </a><p><strong>FIGURE 1</strong>. Conceptual diagrams of methane dynamics in estuarine waters. With no aeration, <strong>(A)</strong> the lowest methane flux comes from a water column that is well oxygenated and <strong>(B)</strong> there is a moderate flux when waters are eutrophic. When aerated, the methane flux is <strong>(C)</strong> highest when there are large bubbles and <strong>(D)</strong> moderate when there are small bubbles.</p></div><div class="clear"></div><div class="DottedLine"></div><p class="mb15">Under this simple conceptual model, it is enticing to speculate that if the bottom waters were re-oxygenated, this would return the aerobic biofilter to its normal state and lower the methane flux to the atmosphere. However, methane is a highly insoluble gas, and the mere addition of air bubbles (devoid of methane) and physical movement of the water during aeration could instead promote methane to dissolve into the rising bubbles and released to the atmosphere. In this case, the size of the air bubbles injected into the bottom waters could affect the magnitude of the atmospheric flux. For example, when the bubbles are large, the physical movement of the turbulent water would release large amounts of methane (<a href="#F1">Figure 1C</a>). When the bubbles are small, there is the possibility for some of the oxygen to diffuse into the surrounding water (which is the goal of aeration systems) and promote aerobic methane oxidation (<a href="#F1">Figure 1D</a>).</p><p class="mb15">Oxygenated water columns are necessary for aerobic methane oxidizing bacteria to help control methane emissions to the atmosphere. In the simplest case of the open ocean where there are deep, well-oxygenated waters, nearly all the dissolved methane in the water column is oxidized (<a href="#B45">Leonte et al., 2017</a>; <a href="#B65">Pohlman et al., 2017</a>). Yet even in these systems, studies have shown there is a lag time for aerobic oxidation (<a href="#B12">Chan et al., 2019</a>), with notable exceptions such as the rapid response of methanotrophs to methane released during events like the Deepwater Horizon oil spill (<a href="#B38">Kessler et al., 2011</a>). Classic works have shown the highest rates of aerobic methane oxidation at the oxycline in arctic lakes (<a href="#B70">Rudd and Hamilton, 1978</a>) suggesting that these organisms are facultative microaerophiles who work in these strong oxygen gradients (<a href="#B63">Oswald et al., 2015</a>; <a href="#B77">Steinle et al., 2017</a>). Aerobic methane oxidation can also proceed at high oxygen concentrations, especially when nitrogen is available. Under eutrophic conditions, when dissolved inorganic nitrogen concentrations >20 µM (M is used throughout as a symbol for mol L<sup>−1</sup>), aerobic methane oxidation occurs at oxygen levels >31 µM; much higher than found in the oxyclines (<a href="#B71">Sansone and Martens, 1978</a>). Along with nitrogen, micronutrients (especially copper) stimulate activity of methane oxidizing bacteria (<a href="#B75">Semrau et al., 2010</a>). Not surprisingly the dissolved methane concentration in the environment has been shown to influence the rate of water column methane oxidation. For example, in Arctic waters, higher rates of methane oxidation were measured when higher concentrations of methane were available in summer (<a href="#B52">Mau et al., 2013</a>). However, during the Deepwater Horizon oil spill, aerobic methane oxidation rates decreased over time after an initial spike, even though methane concentrations remained high (<a href="#B15">Crespo-Medina et al., 2014</a>).</p><p class="mb15">Here we present a study that quantified the effect of engineered aeration on air-water methane emissions from a eutrophic estuary. We leveraged a unique opportunity to manipulate whole ecosystem dissolved oxygen concentrations using two types of engineered destratification systems, that we distinguish based on the bubble size (large bubble and small bubble aeration). These systems provided an ideal opportunity to test the hypothesis that the physical disturbance introduced to a eutrophic system with bubbles will enhance the atmospheric methane flux (<a href="#F1">Figures 1C,D</a>), regardless of oxygen concentration. We also surmised that the two different destratification systems would result in different methane fluxes to the atmosphere based on their bubble size; where the small bubbles would dissolve before reaching the air-water interface and not act as a transfer mechanism like larger bubbles. Furthermore, since the very idea behind engineered aeration is to add oxygen to the water column, we also hypothesized that aeration could enhance aerobic methane oxidation in the water column, which could act to lower the flux of methane to the atmosphere. To test our hypotheses, we conducted experimental manipulations of the aeration systems and sampled surface and bottom waters with aerators ON and then 1–13 days after the aerators were turned OFF. We hypothesized the source of methane in the water was the sediments, thus we also collected sediment cores at the same time. To put the discrete, experimental time points into a longer term context, we also present unique time-series measurements of methane concentrations in bottom waters across the whole estuary. Ultimately, this study contributes data to the growing literature of methane dynamics in shallow, eutrophic environments.</p><a id="h3" name="h3"></a><h2>2 Materials and Methods</h2><h3 class="pt0">2.1 Field Description and Experimental Setup</h3><p class="mb0">The experiments were carried out in a 353-ha tidal tributary to the Chesapeake Bay found in Anne Arundel County, Maryland (<a href="#F2">Figure 2</a>). Rock Creek’s watershed is 80% residential and 20% forested. Due to the poor water quality over the past few decades (e.g., anoxia and extensive algal blooms), the county installed an engineered destratification system in 1988 to bring dissolved oxygen back to former levels (<a href="#B28">Harris et al., 2015</a>). Herein, we call this the “large bubble aeration” system. Up until 2019, this system was made up of 138 ultra-coarse air diffusers distributed along a pipe that lines the middle creek channel with ∼20 mm bubble size (<a href="#B11">CH2M_Hill, 2011</a>), as described in design specifications reported by <a href="#B17">Dames and Moore (1988)</a>. The goal of the system was to vigorously overturn the water column in order to continuously introduce oxygen <em>via</em> de-stratification into the bottom waters, and the system was run continuously throughout the day. Every year, the aerators are turned on June 1 and turned off October 1, in order to minimize the effects of summertime hypoxia. An early study of this system determined that the zone of aeration influence was ∼74 ha and that bottom waters remained oxic when diffusers were ON, but became anoxic within one tidal cycle when aerators were OFF (<a href="#B28">Harris et al., 2015</a>; <a href="#B29">Harris et al., 2016</a>). In spring 2019, the aeration system was upgraded to fine bubble diffusers to provide more oxygen to the water column, herein referred to as the “small bubble aeration” system. The goal of the diffusers is primarily to overturn the water to allow re-aeration to occur at the water surface, not to add oxygen from the bubbles themselves. There are two 213 m long diffusers emanating from the shore mainline, which provide air at a rate of 180 scfm (standard cubic feet per minute) in a continuous bubble pattern (0.26 scfm ft<sup>−1</sup>) of bubbles 3 mm in diameter. The surface water expression of the new aeration system is shown in <a href="#F2">Figure 2B</a>. With this new system, the county public works turns OFF the aerators every night to reduce neighborhood noise and energy consumption.</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 2</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g002.jpg" name="Figure2" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g002.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g002.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g002.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g002.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g002.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g002.jpg" alt="www.frontiersin.org" id="F2" loading="lazy"> </picture> </a><p><strong>FIGURE 2</strong>. Sample location map for Rock Creek. Red lines show location of the aeration tubes, black dots are water column and sediment station locations, white stars are OsmoSampler locations (very close to black dots), and black cross is location of dock where the benthic lander was deployed. <strong>(A)</strong> Location of Rock Creek (black square) in the northern Chesapeake Bay near Baltimore, Maryland. <strong>(B)</strong> Photo of bubbles breaking the surface in 2019 (photo by Laura Lapham).</p></div><div class="clear"></div><div class="DottedLine"></div><p class="mb0">Over the course of 4 years (2016, 2018, 2019, 2021), the water column and sediments were sampled along the creek both within and outside the aeration zone and with both engineering designs (<a href="#T1">Tables 1</a>, <a href="#T2">2</a>). Stations within the aeration zone included RC1, which is located at the up-creek limit of aeration, and RC2 which is found mid-channel and directly in the aeration zone (<a href="#F2">Figure 2</a>). RC7 is ∼1 km downstream from the end of aeration, still within the zone of influence of aeration and where the creek widens, and RC9b is a background site, close to the mouth of the Patapsco River and outside the zone of influence of aeration (<a href="#F2">Figure 2</a>). These stations were introduced in previous studies (<a href="#B28">Harris et al., 2015</a>). Water depths are between 1.5–3.5 m. In 2021, four upstream stations were added to determine the river influence to the aeration zone (<a href="#T2">Table 2</a>).</p><div class="DottedLine"></div><div class="Imageheaders">TABLE 1</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t001.jpg" name="Table1" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t001.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t001.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t001.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t001.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t001.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t001.jpg" alt="www.frontiersin.org" id="T1" loading="lazy"> </picture> </a><p><strong>TABLE 1</strong>. Station information.</p></div><div class="clear"></div><div class="DottedLine"></div><div class="Imageheaders">TABLE 2</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t002.jpg" name="Table2" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t002.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t002.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t002.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t002.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t002.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t002.jpg" alt="www.frontiersin.org" id="T2" loading="lazy"> </picture> </a><p><strong>TABLE 2</strong>. Overview of manipulation experiments.</p></div><div class="clear"></div><div class="DottedLine"></div><p class="mb0">To study the effects of the aeration, our experimental approach was to sample while the aerators had been on for about 1 month, referred to as the “ON” sampling event, which occurred during the daylight hours. In the evening of the “ON” sampling, the aerators were turned off and waters sampled 1 day later (2016; large bubble aeration), 7 days later (2018; large bubble aeration), 13 days later (2019; small bubble aeration), and 3 days later (2021; small bubble aeration); these are referred to as the “OFF” sampling events (<a href="#T2">Table 2</a>). After completion of the experiments, the aerators were turned back ON for the remainder of the season. By sampling at different time periods after aerators were turned OFF, the experiment addressed the question of the impact aeration had on methane flux from the Rock Creek estuary.</p><h3 class="pt0">2.2 Sampling Description</h3><p class="mb0">During each field campaign, we collected water column hydrographic and chemical profiles, discrete water samples from the surface and bottom depths, and shallow (∼30 cm) sediment cores <em>via</em> small boat. Water column temperature, salinity, and dissolved oxygen levels were recorded with a YSI EXO2 multiparameter sonde. In 2018, a benthic lander was placed at a dock (location shown as black cross in <a href="#F2">Figure 2</a>) at the edge of the aeration zone that included continuous temperature, salinity, and dissolved oxygen sensors (YSI EXO2). Salinity is reported in practical salinity scale which has no units. Wind speeds were determined from a handheld anemometer (Weatherhawk, Windmate WM-200).</p><h4>2.2.1. Discrete Water Column Samples</h4><p class="mb0">Water samples were collected for dissolved methane when the aerators were ON and OFF. Water samples were always taken within 1 m of the same GPS location. The sampling location is ∼3 m off-axis to the aerators to ensure to not entangle the boat anchor. Therefore, when the aerators were ON, sampling never occurred in the bubble plume itself, always several meters away. Water column samples for dissolved methane concentrations were collected using published methods (<a href="#B48">Magen et al., 2014</a>). Briefly, a submersible pump was placed at either 50 cm from the air-water interface, or 1 m from the sediment bottom and dispensed water into 125 mL glass serum vials by overfilling 5 times the vial volume and avoiding bubbles. The vials were then capped with thick butyl rubber septa and crimp sealed with aluminum rings. A 10 mL air (Ultra Zero Air purity, Airgas) headspace was given to the vials and then 0.5 mL 8 M KOH was added to arrest microbial activity during storage. The samples were stored upside down at 4°C until they could be measured back at the laboratory with a headspace equilibration technique. In 2021, water was collected with a slight modification to the method where the headspace equilibration step was conducted <em>in situ</em> and then the headspace physically separated from the water sample so no preservation was needed. Briefly, the submersible pump filled 120 mL into a 140 mL plastic syringe, bubble free. Then, 20 mL air (Ultra Zero Air purity, Airgas) was added and shook for 4 min to equilibrate. The temperature of the water was recorded for solubility calculations. Since this was a modification, we conducted efficiency tests using lab standards prior to the field campaign to verify 100% methane recovery from the method (<a href="#SM1">Supplementary Figure S1</a>). A complimentary water sample was also collected to quantify dissolved inorganic nitrogen (DIN) concentrations using published methods (<a href="#B28">Harris et al., 2015</a>), in all years but 2021.</p><h4>2.2.2. Sediments</h4><p class="mb0">Sediment cores were collected by hand off the side of the boat using a 6.5 cm diameter plexiglass cylinder attached to a pole. The cores were brought back to shore and immediately (within 1 h of sampling) sliced into 3 cm vertical sections and the sediment packed into 50 mL centrifuge tubes and stored at 4°C for later analysis of pore-water sulfate concentrations (<a href="#B42">Lapham et al., 2008b</a>). A separate sample for dissolved methane in the pore-waters was also collected by subcoring each section with a 3 mL cut off plastic syringe and placing the material in a 13.5 mL glass serum vial, capped with butyl rubber septae and preserved with 3 mL 1 M KOH (<a href="#B42">Lapham et al., 2008b</a>). Sediment samples were stored at −20°C until analysis.</p><h4>2.2.3. Air Samples</h4><p class="mb0">At each station, a 140 mL plastic syringe was used to collect an air sample above the sampling site. In 2018, more air samples were collected over time because of opportunistic sampling. Notably, on the ON and OFF days, air samples were collected at all stations at dawn, during the day, and at dusk. Sampling in 2019 was synchronous with water and sediment sampling during the day only. The syringe was upwind of any boat traffic and flushed copiously to provide a clean sample, before any other sampling occurred and potentially contaminated the air. These air samples were stored at 23°C for less than 2 days before they were measured for methane concentrations and stable carbon isotope ratios.</p><h4>2.2.4. Continuous Bottom Water Sampling</h4><p class="mb0">To capture the temporal variability of methane concentrations between sampling campaigns and after the aeration manipulation experiments were completed, bottom water was continuously collected using OsmoSamplers at RC1, RC2, and RC7. OsmoSamplers (<a href="#SM1">Supplementary Figures S2A,B</a>) are osmotically-driven pumps that continuously collect and store water in narrow bore copper capillary tubing (<a href="#B33">Jannasch et al., 2004</a>). They have been used in numerous natural environments to quantify dissolved methane concentrations, including from deep water methane seeps (<a href="#B41">Lapham et al., 2008a</a>; <a href="#B83">Wilson et al., 2015</a>), estuaries (<a href="#B24">Gelesh et al., 2016</a>), high altitude rivers (<a href="#B9">Buser-Young et al., 2021</a>), high latitude wetlands (<a href="#B10">Buser-Young et al., 2022</a>) and arctic lakes (<a href="#B57">McIntosh Marcek et al., 2021</a>). Osmosis in the pumps is created by an osmotic potential between a saturated brine chamber and freshwater chamber separated by semi-permeable membranes; no power is needed and there are no moving parts (<a href="#B79">Theeuwes and Yum, 1976</a>; <a href="#B33">Jannasch et al., 2004</a>). The osmotic pump is then connected to small-bore (0.082 cm inner diameter), long (up to 300 m) copper tubing coil, that is prefilled with freshwater. Water is then continuously drawn from the end of the copper tubing and stored in this tubing over time. The copper material is used so gases (i.e., methane) do not diffuse through it. The pumping rate is positively correlated to the surrounding water temperature and the number of membranes in the pump. For these deployments, two pump speeds were used. “Slow” pumps (8 membranes which pump 1 mL day<sup>−1</sup>) were used for the long term collection of bottom water through the summer to give a temporal resolution of about 5 days (deployed for 9 months). “Fast” pumps (44 membranes which pump 5 mL day<sup>−1</sup>) were used for a temporal high resolution of ∼1 day (deployed for 1 week). For the slow pumps, we assume the sample stream undergoes plug flow; thus, dispersion within the tubing is minimal (<a href="#B33">Jannasch et al., 2004</a>). For the fast pumps, plug flow may not be met, so we limited the deployment to a week. OsmoSamplers only collect water thereby precluding gas from affecting the sampler. The intakes are fitted with a 0.2 µm rhizone filter (<a href="#B74">Seeberg-Elverfeldt et al., 2005</a>) to preclude microbes from the collection and alter the sample stream in the tubing during the deployment.</p><p class="mb0">At the time of the deployment, the OsmoSamplers were attached to the copper coils and placed in a plastic crate (33 × 33 × 28 cm) under 16 kg weight and tied to a surface buoy (<a href="#SM1">Supplementary Figures S2C–E</a>). In 2018, the fast OsmoSampler sets were deployed at stations RC1, RC2, and RC7. In 2018, and 2019, slow OsmoSampler sets were also deployed at each of those stations along with Onset temperature and conductivity loggers. Because Rock Creek is a dynamic estuary, we used the conductivity detectors to verify the time-stamps in the OsmoSampler coils. OsmoSamplers were deployed from 9 July to 9 October 2018 and 18 June 2019 to 28 October 2019.</p><p class="mb0">Upon recovery, the copper coils were sealed on either end with pliers and taken back to the lab to be stored at 4°C prior to further processing in the lab. The sensor data were downloaded. Within 1 week, the copper coils were unspooled and crimped into alternating lengths of 50 cm and 4.5 m using a wire crimping tool (<a href="#B24">Gelesh et al., 2016</a>). The 50 cm sections were squeezed with a bench-top roller to flatten the copper tubing and force the liquid into 2 mL plastic tubes to immediately test for salinity using a handheld Extech RF20 refractometer. Because the coils were prefilled with freshwater before deployment, sectioning was terminated when zero salinity was observed for three samples in a row. These samples were also measured for chloride concentrations to compare to sensor conductivity measurements to verify time-stamps (<a href="#B24">Gelesh et al., 2016</a>). The 4.5 m copper coil sections were squeezed using the bench-top hand roller which expressed sample liquid through a gastight adaptor and needle, and into a 13.5 mL glass sample vial at the opposite end, previously capped with a butyl rubber septum to prevent gas exchange, and flushed with helium. Each 4.5 m copper section contained approximately 2 mL of sample liquid that was transferred to the vials, resulting in an initial overpressure of approximately 2 mL. The dissolved CH<sub>4</sub> equilibrated with the helium headspace after shaking the vial for 2 min. Time stamps were calculated by adjusting pumping rates to <em>in situ</em> temperature, as shown in <a href="#B24">Gelesh et al. (2016)</a>. Unfortunately, the fast pumps only had 1 weeks’ worth of tubing yet were deployed for 12 days due to weather delays. Thus, the pumps overpumped the coil and the first part of the deployment was lost.</p><h3 class="pt0">2.3 Analytical Methods</h3><p class="mb0">For water column and sediment vial samples, the headspace was equilibrated with the dissolved methane from the aqueous sample and the headspace analyzed for the ppmv methane. For all samples, an aliquot of the headspace was extracted from the vials and injected onto a gas chromatograph (SRI 8610C multi-gas) equipped with a HayeSep D packed column and a Flame Ionization Detector in order to quantify methane concentrations (<a href="#B48">Magen et al., 2014</a>). Certified standards (Airgas, Inc.) were used for the calibration curve. Analytical precision is 3% and all measurements were above detection limit of 2 ppmv. Resultant partial pressures were then used to calculate dissolved methane concentrations (in nM) in either the water column or the porewater using Henry’s law according to equations in <a href="#B48">Magen et al. (2014)</a> and porosity corrections according to <a href="#B41">Lapham et al. (2008a)</a>.</p><p class="mb0">For sulfate and chloride concentrations in the sediment porewater samples, the tube containing whole sediment was centrifuged (3000 RPM, 30 min, 20°C, Sorvall<sup>©</sup> RT 6000D) and the resultant supernatant filtered with a 0.2 µm syringe filter. Samples were then diluted (1:135) in Milli-Q water and analyzed on a Dionex ICS 1000 ion chromatograph (IonPac AG22 4 × 50 mm guard column, IonPac AS22 4 × 250 mm analytical column, and ASRS 300 4 mm suppressor) with an AS40 Autosampler. Water samples from the 50 cm OsmoSampler sections were also measured for chloride with the same dilution to calculate salinity. Certified IAPSO seawater standard (Ocean Scientific International Ltd.) was used for the calibration curve. Analytical precision is 2% and all measurements were above detection limit of 0.05 mM for sulfate.</p><p class="mb0">Air syringes were directly connected to the intake of a cavity ring down spectrometer (CRDS, Picarro 2201i) to measure for methane concentrations and methane stable carbon isotopes. For the water samples, 10 mL of degassed brine was added to the vials to displace the headspace and injected into the small sample isotope module (Picarro, Inc.) to introduce a small sample to the CRDS, similar to procedure in <a href="#B57">McIntosh Marcek et al. (2021)</a>. Isotope values were obtained through calibration with three Vienna Pee Dee Belemnite (VPDB) referenced standards (−23.9‰, −38.3‰, and −66.5‰ (±0.2‰); Isometric Instruments). Isotopic results are reported using the δ<sup>13</sup>C notation in per mil (‰), where δ<sup>13</sup>C = (R<sub>sample</sub>/R<sub>standard</sub> -1)*1,000 and R = <sup>13</sup>C/<sup>12</sup>C. Analytical precision is 2% for concentrations and 4‰ for stable carbon isotope ratios.</p><h3 class="pt0">2.4 Calculations</h3><p class="mb0">The air-water flux of CH<sub>4</sub>, F, was determined for all discrete sampling campaigns using the updated flux equations presented <a href="#B80">Wanninkhof (2014)</a>:</p><div class="equationImageholder"><math id="e1"><mrow><mi>F</mi><mo>=</mo><mi>k</mi><mrow><mo>(</mo><mrow><msub><mi>C</mi><mi>w</mi></msub><mo>−</mo><msub><mi>C</mi><mrow><mi>e</mi><mi>q</mi></mrow></msub></mrow><mo>)</mo></mrow></mrow><mspace width="5em"></mspace><mo stretchy='false'>(</mo><mn>1</mn><mo stretchy='false'>)</mo></math><div class="clear"></div></div><p class="noindent">where k is the gas transfer velocity (length time<sup>−1</sup>), C<sub>w</sub> is the measured surface water concentration, and C<sub>eq</sub> is the CH<sub>4</sub> concentration in equilibrium with the atmosphere at <em>in situ</em> conditions (<a href="#B84">Yamamoto et al., 1976</a>). There are several versions of <a href="#e1">Eq. 1</a>, mostly based on wind speed. Here we employ the formulation and parameterization of <a href="#B61">Myllykangas et al. (2020)</a>, which reports a k value adapted from <a href="#B66">Raymond and Cole (2001)</a>:</p><div class="equationImageholder"><math id="e2"><mrow><mi>k</mi><mo>=</mo><mn>1.91</mn><mo> </mo><msup><mi>e</mi><mrow><mn>0.35</mn><mi>u</mi></mrow></msup><msup><mrow><mrow><mo>(</mo><mrow><mfrac><mrow><mi>S</mi><mi>c</mi></mrow><mrow><mn>600</mn></mrow></mfrac></mrow><mo>)</mo></mrow></mrow><mrow><mo>−</mo><mn>0.5</mn></mrow></msup></mrow><mspace width="5em"></mspace><mo stretchy='false'>(</mo><mn>2</mn><mo stretchy='false'>)</mo></math><div class="clear"></div></div><p class="noindent">where <em>u</em> is the average wind speed and Sc is the Schmidt number for CH<sub>4</sub> in freshwater calculated from <a href="#B80">Wanninkhof (2014)</a>. Wind speeds were obtained from a nearby NOAA buoy (National Data Buoy Center BLTM2, Baltimore, MD) and averaged over the 3 days prior to sampling. Since the buoy is located 15 km away from Rock Creek, we compared the buoy to the handheld anemometer readings and found they compared within 7%. The buoy wind speed was used for all stations. Wind speeds varied between 2–3.2 m s<sup>−1</sup>, which translated to k values varying between 3.8 and 5.9 cm h<sup>−1</sup>, similar to values found in <a href="#B47">MacIntyre et al. (2010)</a> in a lake system and mangrove dominated estuaries (<a href="#B69">Rosentreter et al., 2017</a>). Air-water fluxes were then calculated using <a href="#e1">Eq. 1</a> and reported as µmol CH<sub>4</sub> m<sup>−2</sup> d<sup>−1</sup>.</p><p class="mb0">Since the calculated air-water flux is inherently based on assumptions of a stagnant boundary layer, the calculated air-water methane flux when the aerators are ON will be underestimated. To constrain this better when the aerators were ON, we calculated an air-water methane flux from direct bubble transport to surface water by applying the bubble radius of the system (3 mm, Mobley Engineering, Inc., personal communication) to an existing bubble model output to determine the mass transfer coefficient (<a href="#F4">Figure 4</a> in <a href="#B56">McGinnis and Little, 2002</a>). This mass transfer coefficient for the bubble radius in the system is 0.04 cm s<sup>−1</sup> (or 144 cm h<sup>−1</sup>) which was then used in <a href="#e1">Eq. 1</a> to calculate a modified air-water methane flux for 2018 and 2019 at RC1 and RC2.</p><p class="mb0">The sediment-water methane diffusive flux was calculated from Fick’s first law:</p><div class="equationImageholder"><math id="e3"><mrow><msub><mi>J</mi><mrow><mi>C</mi><mi>H</mi><mn>4</mn><mo>−</mo><mi>S</mi><mi>W</mi><mi>I</mi></mrow></msub><mo>=</mo><mo> </mo><mo>−</mo><mi>φ</mi><msub><mi>D</mi><mi>s</mi></msub><mfrac><mrow><mi>d</mi><mi>C</mi></mrow><mrow><mi>d</mi><mi>x</mi></mrow></mfrac></mrow><mspace width="5em"></mspace><mo stretchy='false'>(</mo><mn>3</mn><mo stretchy='false'>)</mo></math><div class="clear"></div></div><p class="noindent">where J<sub>CH4-SWI</sub> is the methane flux (µmol CH<sub>4</sub> cm<sup>−2</sup> yr<sup>−1</sup>) at the sediment-water interface, φ is the porosity (0.8), D<sub>s</sub> is the sedimentary methane diffusion coefficient (cm<sup>2</sup> s<sup>−1</sup>), x is the vertical sediment depth (cm) and dC/dx is the concentration gradient of methane. D<sub>s</sub> was calculated for each station, corrected for tortuosity and <em>in situ</em> pressures (based on water depth), temperatures, and salinity (<a href="#B59">Millero, 1996</a>), and was ∼1.7 × 10<sup>–6</sup> cm<sup>2</sup> s<sup>−1</sup>. The gradient term was calculated between the uppermost porewater measurements; which usually started at 1.5 cm into the sediment. Thus, this gradient is most likely overestimated because it ignores any oxidation processes that might occur in that upper 1.5 cm of sediment. Using this gradient, the diffusive fluxes were then calculated for each station and time point with <a href="#e3">Eq. 3</a>. For convenience, fluxes are reported as positive but represent flux out of sediment).</p><h3 class="pt0">2.5 Box Model</h3><p class="mb0">We applied a simple box model to the flux data with the assumption that the sediments are the only source of methane to the water, and the atmospheric flux was the only sink of methane in the water. If the two balanced, then there would be no additional contributions to the methane budget in Rock Creek. If the fluxes did not balance, we could invoke additional microbial oxidation or production in the water column and/or advective transport of methane from up- or down-stream. To do this, at each station for 2018 and 2019, we subtracted the atmospheric methane flux from the sedimentary methane flux, then assigned the net difference as the water column methane inventory anomaly. For the aeration sites (RC1 and RC2), both air-water and sedimentary fluxes were averaged together. We should note that when the aerators are ON, we used the fluxes estimated from the stagnant boundary layer model which most likely underestimates the atmospheric flux.</p><h3 class="pt0">2.6 Statistics</h3><p class="mb0">Student t-test was used in Excel (two-tailed, paired) to determine if the methane concentrations calculated from the discrete sampling events and the OsmoSampler samples were significantly similar. To create the paired methane dataset for this test, we extracted the OsmoSampler methane concentration that was sampled at the same time as the discrete methane measurement from the water column at each station. Please note that the OsmoSamplers average over 6–12 h. This comparison was done with the fast OsmoSamplers during 2018 (see Section 3.6 for result).</p><a id="h4" name="h4"></a><h2>3 Results</h2><h3 class="pt0">3.1 Water Column Oxygen, Dissolved Inorganic Nitrogen, Temperature, and Salinity</h3><p class="mb0">The experimental design was intended to measure methane concentrations in well oxygenated, bubble-influenced waters when aerators were ON, and then hypoxic or anoxic waters without bubble transport when the aerators were turned OFF. Based on previous experiments in Rock Creek, the change to hypoxic conditions occurred within 1 day of turning off the aerators (<a href="#B28">Harris et al., 2015</a>). However, for our manipulations, the goal was to observe changes over the longer term (up to 13 days) and the systems response. It is important to remember there was a 7-day difference between ON and OFF treatments in 2018 and 13-day difference in 2019.</p><p class="mb0">In 2018 during aeration, the dissolved oxygen (DO) concentrations were ∼7 mg L<sup>−1</sup> at RC1 and RC2, and well mixed (<a href="#F3">Figure 3</a> top panel). At RC7, the water column was still oxygenated (>4 mg L<sup>−1</sup>; comparable to conditions at RC9b). After aerators were OFF for 7 days, RC1 and RC2 became hypoxic throughout the water column below 0.5 m, and the other stations had weakly stratified water columns, but with hypoxic bottom waters below 2.5 m. Dissolved inorganic nitrogen concentrations averaged 14.4 µM (ranging between 0.8 and 19 µM for all stations) when aerators were ON and 7.2 µM (ranging between 3.7 and 12 µM) when aerators were OFF (<a href="#T2">Table 2</a>). Salinity was ∼5–5.5 (<a href="#F3">Figure 3</a>). Water temperatures in 2018 varied between 26–30°C, and were similar for both ON and OFF treatments (data not shown).</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 3</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g003.jpg" name="Figure3" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g003.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g003.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g003.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g003.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g003.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g003.jpg" alt="www.frontiersin.org" id="F3" loading="lazy"> </picture> </a><p><strong>FIGURE 3</strong>. Water column salinity and dissolved oxygen (O<sub>2</sub>) profiles from 2018 (top panels) and 2019 (bottom panels) for stations within the aeration zone and stations outside the aeration zone. Blue symbols signify when aerators were ON and red when they were OFF.</p></div><div class="clear"></div><div class="DottedLine"></div><p class="mb0">In 2019, dissolved oxygen was relatively high in surface waters when the aerators were OFF at RC1 and RC2, giving way to oxygen-depleted conditions below 2 m (<a href="#F3">Figure 3</a>). Dissolved inorganic nitrogen concentrations averaged 27.8 µM (ranging between 5 and 36 µM for all stations) when aerators were ON and 4.9 µM (ranging between 1 and 13 µM) when aerators were OFF (<a href="#T2">Table 2</a>). Salinity was lower in 2019 than in 2018 and was nearly 3.5 when aerators were ON, and 4.5 when aerators were OFF (<a href="#F3">Figure 3</a>). Water temperatures were between 26 and 29°C during the ON treatment and were warmer during OFF treatment (data not shown).</p><p class="mb0">The tidal stage at each station varied over the discrete sampling time points (<a href="#SM1">Supplementary Figure S3</a>). In 2018 when aerators were ON, RC7 was sampled first at the ebbing tide, RC9b at low tide, and RC1 and RC2 at a high tide. When aerators were OFF, RC1 and RC2 were collected close to high tide or when waters were just beginning to ebb. RC7 and RC9b were sampled on the ebb tide. In 2019 when aerators were ON, RC9b, RC1, and RC2 were collected on flooding tide, and RC7 was collected right after high tide. When aerators were OFF, all stations were collected near the high tide.</p><h3 class="pt0">3.2 Methane Concentrations and Stable Carbon Isotope Ratios in Water</h3><p class="mb0">Dissolved methane concentrations in surface and bottom water of Rock Creek varied over space and time (<a href="#F4">Figure 4</a>). Overall, concentrations ranged between 150 and 1,500 nM, orders of magnitude higher than atmospheric equilibrium (which is ∼3 nM), and were higher at stations RC1 and RC2, within the aeration zone, compared to stations closer to the Patapsco River (RC7, RC9b). In 2016, concentrations at RC2 and RC7 were around 400 nM throughout the period of measurements, regardless of aeration status (<a href="#F4">Figure 4A</a>). In 2018, during aeration, the waters were relatively well mixed between surface and bottom waters (<a href="#F4">Figure 4B</a>). Once aerators were turned OFF, there was an increase in bottom water methane at RC1 and RC2, and not much change at RC7 and RC9b. After 7 days, the surface waters were enriched in methane compared to the bottom water. In 2019, the concentrations between surface and bottom water followed expectations: during aeration, the water column was well mixed so there was little difference between surface and bottom waters and when aerators were turned OFF, methane concentrations were higher in the bottom water than surface waters after 13 days (<a href="#F4">Figure 4C</a>).</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 4</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g004.jpg" name="Figure4" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g004.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g004.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g004.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g004.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g004.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g004.jpg" alt="www.frontiersin.org" id="F4" loading="lazy"> </picture> </a><p><strong>FIGURE 4</strong>. Methane concentrations (colored bars) in surface (S) and bottom (B) water at all stations in <strong>(A)</strong> 2016, <strong>(B)</strong> 2018, <strong>(C)</strong> 2019, and <strong>(D)</strong> 2021. All blue colors represent the “ON” situation, and the gradients in gray color represent the number of days after aerators were turned off, which are described in each panel. Error bars represent standard error on replicate samples collected.</p></div><div class="clear"></div><div class="DottedLine"></div><p class="mb0">The 2021 field campaign was designed to resolve the upstream contribution of methane to the aeration zone. Overall, methane concentrations were lower than 2019 values but also showed the same pattern of higher concentrations at RC1 and RC2, than at RC7 (<a href="#F4">Figure 4D</a>). At RC7, concentrations were similar between surface and bottom, and regardless of aeration status. At RC1 and RC2, methane concentrations were lower when aerators were OFF, which was unexpected. The other unexpected result was to record higher methane concentrations in surface waters than bottom waters during both ON and OFF periods (<a href="#F4">Figure 4D</a>). Measurements made upstream of the aerators, which were only made in 2021, showed that surface waters were always higher than the bottom waters and that concentrations during the ON treatment were always higher than the OFF. Furthermore, methane concentrations were highest in the most upstream station, and declined downstream and into the estuary, such that water flowing into the aeration zone from upstream were enriched with methane relative to the aeration zone itself.</p><p class="mb0">The water column methane concentrations are also presented in <a href="#F5">Figures 5A–D</a>, <a href="#F6">6A–D</a> as compilation figures showing the water column and sediments in a holistic view. Here we add to the water column concentration data the stable isotopic ratio of methane carbon to distinguish source of this methane (<a href="#F5">Figures 5I–L</a>, <a href="#F6">6I–L</a>). Regardless of station or aeration status, δ<sup>13</sup>C-CH<sub>4</sub> values ranged from −69 to −51‰, with an average value of −61.4 ± 3.6‰, which is near the standard deviation of the method.</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 5</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g005.jpg" name="Figure5" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g005.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g005.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g005.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g005.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g005.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g005.jpg" alt="www.frontiersin.org" id="F5" loading="lazy"> </picture> </a><p><strong>FIGURE 5</strong>. In 2018, methane concentrations <strong>(A–H)</strong> and stable carbon isotopes <strong>(I–O)</strong> in water column (blue background) and in sediments (brown background). Blue symbol color denotes when aerators were on, and dark gray when they were off for 7 days. Horizontal bars in sediments <strong>(M–O)</strong> shows the movement of the sulfate-methane transition zone between ON (blue color) and OFF (gray color).</p></div><div class="clear"></div><div class="DottedLine"></div><div class="Imageheaders">FIGURE 6</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g006.jpg" name="Figure6" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g006.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g006.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g006.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g006.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g006.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g006.jpg" alt="www.frontiersin.org" id="F6" loading="lazy"> </picture> </a><p><strong>FIGURE 6</strong>. In 2019, methane concentrations <strong>(A–H)</strong> and stable carbon isotopes <strong>(I–O)</strong> in water column (blue background) and in sediments (brown background). Blue symbol color denotes when aerators were on, and dark gray when they were off for 13 days. Horizontal bars in sediments <strong>(M–O)</strong> shows the movement of the sulfate-methane transition zone between ON (blue color) and OFF (gray color).</p></div><div class="clear"></div><div class="DottedLine"></div><h3 class="pt0">3.3 Sediment Porewater</h3><p class="mb0">Methane concentrations measured from the sediment porewaters increased with sediment depth (<a href="#F5">Figures 5</a>, <a href="#F6">6</a>, brown colored panels). Surface concentrations were at ∼µM levels and increased to as high as 1,300 µM at the bottom of the core. In both years, concentrations were higher in the aeration zone (RC1 and RC2) and outside the aeration zone (RC7) compared to the background site (RC9b). Yet, there is variability in the sediment profiles. For example, at RC1 and RC2, sediment methane concentrations were lower in 2018 than in 2019, regardless of aeration status. After 13 days of no aeration, methane concentrations were higher in the sediments at RC2 and RC7. As noted, <a href="#F5">Figures 5</a>, <a href="#F6">6</a> also contain methane water column parameters for comparison purposes.</p><p class="mb0">Methane increases in sediment porewaters are typically associated with a drawdown of sulfate in the surficial depths due to sulfate reducers outcompeting methanogens for substrates (<a href="#B31">Hoehler et al., 1994</a>). Therefore, we also measured sulfate in porewater to help our understanding of anaerobic biogeochemical processing. Sulfate concentrations decreased downcore in all stations except RC9b although the depth of low sulfate (SO<sub>4</sub> < 0.5 mM) varied (<a href="#SM1">Supplementary Figure S4</a>). The depth of low sulfate typically coincides with the increase of methane, and is known as the sulfate methane transition (SMT) depth. The SMT is an area of active anaerobic methane oxidation <em>via</em> sulfate reduction (<a href="#B36">Jørgensen et al., 2020</a>) and is a useful metric to show how active the anaerobic microbial community is in a sediment column (<a href="#F5">Figures 5M–O</a>, <a href="#F6">6M–O</a>). While we expected to see the SMT depth shoal when aeration was turned OFF, there was no consistent pattern of the depth of the SMT with aeration status for both years, although we will specifically present SMT depth patterns in each year below. Using the gradients from the top of the cores, the methane flux to the sediment-water interface varied across space and time, and ranged between 0.1 and 700 µmol m<sup>−2</sup> d<sup>−1</sup> (<a href="#SM1">Supplementary Figure S5</a>).</p><p class="mb0">Chloride concentrations were also measured as a conservative tracer and as a way to validate any depletion of sulfate coming from sulfate reduction and not a consequence of groundwater. In 2018, while the chloride concentrations showed a slight increase in depth, the depth averages are as follows: 76 ± 8 mM (RC1, RC2), 100 ± 14 mM (RC7), and 85 ± 10 mM (RC9). Since these chloride values are within the range of what would be expected given the overlying water salinity, we represent any conservative mixing in terms of how it might affect sulfate concentrations. Given the rule of constant proportions, we calculated the range of sulfate values that would be estimated given those chloride concentrations (shaded rectangles in <a href="#SM1">Supplementary Figure S4</a>).</p><p class="mb0">Methane stable carbon isotope ratios were measured to help distinguish the fate of methane formed in the sediment. Our measurements revealed two patterns in δ<sup>13</sup>C-CH<sub>4</sub> values with depth: 1) δ<sup>13</sup>C-CH<sub>4</sub> increasing with depth and 2) δ<sup>13</sup>C-CH<sub>4</sub> peaks at intermediate depths associated with the SMT. In 2018, at RC1, the δ<sup>13</sup>C-CH<sub>4</sub> values were similar during the ON and OFF conditions in that the surface was <sup>13</sup>C depleted (between −80 and −70‰), they became heavier with depth to as high as −50‰, and then decreased again to near surface sediment values (<a href="#F5">Figure 5M</a>). The depth of the SMT deepened with aerators OFF. At RC2, during the ON condition, δ<sup>13</sup>C-CH<sub>4</sub> values were ∼−80‰ and increased with depth in the core (<a href="#F5">Figure 5N</a>). During the OFF condition, δ<sup>13</sup>C-CH<sub>4</sub> values were around −60‰ in the shallow depths and quickly decreased to −80‰ at the SMT. At RC7, just outside the aeration zone, δ<sup>13</sup>C-CH<sub>4</sub> values were ∼−70‰ at the surface when waters were aerated and decreased to −85‰ at the bottom of the core (<a href="#F5">Figure 5O</a>). During the OFF situation, values showed a similar trend at the surface but then showed a mid-depth minimum of −80‰ at the SMT.</p><p class="mb0">In 2019, the δ<sup>13</sup>C-CH<sub>4</sub> value trends showed more consistency across the stations (<a href="#F6">Figures 6M–O</a>). At RC1, values decreased at the surface from as high as −40 to ∼ −70‰ where values remained for about 15 cm into the sediments (<a href="#F6">Figure 6M</a>). RC2 and RC7 show almost the same isotope profiles where values decrease downcore, but the δ<sup>13</sup>C-CH<sub>4</sub> values are ∼10–15‰ higher when aerators were ON compared to when they were OFF (<a href="#F6">Figures 6N,O</a>).</p><h3 class="pt0">3.4 Methane Concentrations and Stable Carbon Isotope Ratios in Air and Air-Water Fluxes</h3><p class="mb0">Methane concentrations were measured in the air above each station during 2018 and 2019. In 2018, the average atmospheric methane concentration across all sites was 1.84 ± 0.06 ppmv, and didn’t vary between ON and OFF conditions, except at the dawn sampling (<a href="#F7">Figures 7A,B</a>). The average δ<sup>13</sup>C-CH<sub>4</sub> value of the background methane was −51.4 ± 10‰ (<a href="#F7">Figures 7C,D</a>). Dawn sampling on the ON and OFF days showed elevated methane concentrations, reaching as high as 3 ppmv at RC7 which had a δ<sup>13</sup>C-CH<sub>4</sub> value of −90‰ (<a href="#F7">Figures 7A,B</a>).</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 7</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g007.jpg" name="Figure7" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g007.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g007.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g007.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g007.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g007.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g007.jpg" alt="www.frontiersin.org" id="F7" loading="lazy"> </picture> </a><p><strong>FIGURE 7</strong>. Methane concentrations in air above the water in <strong>(A)</strong> 2018 and <strong>(B)</strong> 2019. δ<sup>13</sup>C-CH<sub>4</sub> values in air above the water in <strong>(C)</strong> 2018 and <strong>(D)</strong> 2019. Shaded regions indicates when aerators were OFF.</p></div><div class="clear"></div><div class="DottedLine"></div><p class="mb0">To calculate the air-water methane flux, we used two approaches. The first used the stagnant boundary layer model and most likely underestimates the flux for when the aerators are ON. Using this model, the air-water methane fluxes ranged between 300 and 1,500 µmol CH<sub>4</sub> m<sup>−2</sup> d<sup>−1</sup> (<a href="#F8">Figure 8</a>). The flux was higher at RC1 and RC2 than other stations, regardless of aeration status or year. In 2018, the flux at the RC1 and RC2 was higher when aerators were ON after 7 days, whereas in 2019, the flux was lower when the aerators were ON. The second approach was only carried out when the aerators were ON and assumed methane was being stripped from the water as the aerator bubbles traveled up the water column. The calculated fluxes were much higher than the fluxes from the stagnant boundary layer (<a href="#F8">Figure 8</a> extended arrows to blue dots). In 2018, at RC1 and RC2, air-water methane flux was 30,730 and 19,380 µmol CH<sub>4</sub> m<sup>−2</sup> d<sup>−1</sup>, respectively. In 2019, at RC1 and RC2, air-water methane flux was 14,669 and 8,342 µmol CH<sub>4</sub> m<sup>−2</sup> d<sup>−1</sup>, respectively.</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 8</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g008.jpg" name="Figure8" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g008.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g008.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g008.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g008.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g008.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g008.jpg" alt="www.frontiersin.org" id="F8" loading="lazy"> </picture> </a><p><strong>FIGURE 8</strong>. Air-water methane flux for 2018 and 2019. Aerated waters are in blue, and non-aerated waters are shown in gray scale that corresponds to the number of days aerators were off. The bars indicate the flux calculated with the stagnant boundary layer model, whereas the extended arrows to the blue dots indicate the flux recalculated with bubble influence.</p></div><div class="clear"></div><div class="DottedLine"></div><h3 class="pt0">3.5 Box Model Results</h3><p class="mb0">The sediment and air-water fluxes were used in the box model to determine if there are additional sources or sinks of methane beyond what is coming from the sediments and being lost to the atmosphere (<a href="#F9">Figure 9</a>). In 2018, during large bubble aeration, there was a large source of methane (positive values in <a href="#F9">Figure 9</a>) at the aerators, and actually a methane sink from RC7 when the aerators were off. In 2019, there was a methane source across all stations, regardless of aeration status, and this source was fairly constant across the sites (<a href="#F9">Figure 9</a>). The exception to this was at the aerators when they were ON; there was a large methane sink (negative value in <a href="#F9">Figure 9</a>). However, since the atmospheric flux during the ON status is mostly likely underestimated, this exception is most likely a methane source too.</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 9</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g009.jpg" name="Figure9" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g009.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g009.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g009.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g009.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g009.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g009.jpg" alt="www.frontiersin.org" id="F9" loading="lazy"> </picture> </a><p><strong>FIGURE 9</strong>. Water column methane inventory anomaly assuming the sediments are the only source of methane to the water and the air-water interface is the only sink. A positive value means that there must be a source of methane to balance the source and sink, whereas the negative value means there must be a sink consuming methane.</p></div><div class="clear"></div><div class="DottedLine"></div><h3 class="pt0">3.6 Time-Series Water Column Methane Measurements</h3><p class="mb0">Using OsmoSamplers, two separate records of dissolved methane concentrations from bottom water were obtained. The first was from the OsmoSampler deployment in 2018 that spanned 1-week using fast pumps with ∼1 day resolution (<a href="#F10">Figure 10</a>). The time stamps assigned were verified by comparing salinity (as calculated from chloride concentrations) in the OsmoSampler coils and the salinity from the sensor packages (<a href="#SM1">Supplementary Figure S6</a>). The salinity comparison shows relatively good agreement, especially at station RC7. The overall trend is similar between the sensor and OsmoSamplers at RC1 and RC2, but as we have observed in previous studies, the absolute salinity values did not match well at these stations (<a href="#B24">Gelesh et al., 2016</a>). The highest methane concentrations came 3 days after the aerators were turned OFF at RC2 and reached almost 3,000 nM (<a href="#F10">Figure 10</a>). Concentrations were also high at RC1 during this time. The timing of this methane peak came right after an event where dissolved oxygen (measured between stations RC2, and RC7) increased rapidly to ∼8 mg L<sup>−1</sup> oxygen (<a href="#F10">Figure 10</a>). After 14 July 2018, methane concentrations decreased to less than 1,000 nM and were similar at all stations. Methane concentrations from OsmoSamplers were cross-checked with our discrete samples and we see no statistical difference between the two (<em>p</em> = 0.72); which is the first time this has been verified in field tests. Methane concentrations at RC7 remained lower than the other stations.</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 10</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g010.jpg" name="Figure10" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g010.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g010.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g010.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g010.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g010.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g010.jpg" alt="www.frontiersin.org" id="F10" loading="lazy"> </picture> </a><p><strong>FIGURE 10</strong>. Methane concentrations in high temporal resolution in bottom water from fast OsmoSamplers (black and white symbols) and discrete water samples (red symbols) in 2018 from stations RC1 (filled stars), RC2 (open stars), and RC7 (filled circles). Discrete samples overlap with OsmoSampler concentrations. Thin black line shows a dissolved oxygen record from sensors deployed in the bottom water off a nearby dock.</p></div><div class="clear"></div><div class="DottedLine"></div><p class="mb0">The second time-series record of methane concentration came from OsmoSamplers deployed through the summer and into the fall of 2018 and 2019, and contained slow pumps that give ∼ weekly resolution (<a href="#F11">Figure 11</a>). The temporal pattern was not the same each year. In 2018, at RC1 and RC2, the initial concentrations before aeration were lower (∼400 nM), and then almost doubled when aerators were turned OFF (<a href="#F11">Figure 11A</a>). Once they were turned back ON after our experiment, concentrations at RC1 continued to decrease at a rate of ∼13 nM day<sup>−1</sup> (linear fit with <em>R</em><sup>2</sup> = 0.8); whereas at RC2, concentrations continued to increase through July and finally peak in August at 17,000 nM. Concentrations at RC2 then decreased and reached ∼1,000 nM for the remainder of the timeseries. At RC7, concentrations didn’t show much change with time and averaged 842 ± 265 nM. In 2019, we captured much higher temporal resolution with the samplers which started about 2 weeks before our experiment began (<a href="#F11">Figure 11B</a>). Overall, concentrations were lower than in 2018 and ranged between 110 and 1,667 nM. There were concentration differences across sites, where methane concentrations at RC1 averaged 368 ± 100 nM; RC2 averaged 558 ± 136 nM; and RC7 averaged 400 ± 270 nM (<a href="#F11">Figure 11B</a>). The bottom water temperature varied between 23–28°C in 2018 with some variability (<a href="#SM1">Supplmentary Figure S7</a>), whereas in 2019, the temperature gradually increased from ∼23°C to a high of ∼30°C in August and then decreased into the fall where a sudden decreased to less than 20°C occurred when the aerators were turned off (<a href="#SM1">Supplementary Figure S7</a>).</p><div class="DottedLine"></div><div class="Imageheaders">FIGURE 11</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g011.jpg" name="Figure11" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g011.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g011.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g011.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g011.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g011.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-g011.jpg" alt="www.frontiersin.org" id="F11" loading="lazy"> </picture> </a><p><strong>FIGURE 11</strong>. Methane concentrations in bottom water in <strong>(A)</strong> 2018, and <strong>(B)</strong> 2019. Shaded region indicates when the aerators were turned off, otherwise, they were on.</p></div><div class="clear"></div><div class="DottedLine"></div><a id="h5" name="h5"></a><h2>4 Discussion</h2><p class="mb15">One solution to estuarine eutrophication is to artificially aerate the waters with bubble systems. This solution has benefits for reintroducing oxygen back into the water, but it could also have consequences for methane cycling. Previous studies have documented that the hypoxic or anoxic conditions in bottom waters that result from eutrophication also lead to the build-up of dissolved methane diffusing into the bottom waters from the sediments (<a href="#B3">Bange et al., 2010</a>; <a href="#B24">Gelesh et al., 2016</a>) which can result in a greater atmospheric flux. Thus, we hypothesized that when aerators are placed in such a system, the physical movement of all that water, with the fact that methane has low solubility, would enhance an atmospheric methane flux. To our knowledge, there is only one other study in a temperate lake that has studied oxygen effects on methane dynamics, and they found a remarkable decrease in methane build-up with engineered aeration (<a href="#B32">Hounshell et al., 2021</a>), yet they did not quantify air-water flux. With our dataset, we were able to directly calculate this flux when the aerators were ON versus OFF to determine the impact of aeration in terms of methane dynamics. In addition to this, we also considered that the addition of oxygen to the water column might stimulate microbial aerobic methane oxidation which would somewhat control the release of methane at the air-water interface. Through our whole ecosystem manipulation experiment, we were able to address the following questions: 1) what is the effect of aeration on the atmospheric methane flux, 2) is Rock Creek an atmospheric methane source, 3) what is the source of water column methane in the Rock Creek, and 4) is aerobic methane oxidation enhanced in the water column? We also gained insights into complex biogeochemical processes and potential feedbacks occurring in this sub-estuary during and after aeration that sharpens our focus for future studies to further elucidate critical mechanisms related to dissolved oxygen dynamics and associated biogeochemical effects.</p><h3 class="pt0">4.1 Aeration Enhanced Atmospheric Methane Flux</h3><p class="mb0">We hypothesized that the air-water methane flux would be higher during aeration then when the aerators were OFF. When we simply apply the stagnant boundary layer model to calculate the fluxes, we see that sites within the aeration zone (RC1 and RC2) had higher methane fluxes than downstream, regardless of aeration status (<a href="#F8">Figure 8</a>). Yet these fluxes are most likely underestimates when the aerators are ON, as we see with the modified flux calculation (<a href="#F8">Figure 8</a>, extended arrows to blue dots). We further hypothesized that under small bubble aeration, the flux would be lower than under large bubble aeration (<a href="#F1">Figure 1</a>); which is what the fluxes showed in 2018 (large bubble aeration) versus 2019 (fine bubble aeration, <a href="#F8">Figure 8</a>), supporting this hypothesis. The other observation was that in 2019, when the fine bubble aeration was installed, the air-water methane flux was lower during aeration then when the aerators were turned OFF. We have already stated that this flux is most likely underestimated. Future work would benefit from directly measuring this flux with floating chambers to more precisely quantify this flux.</p><h3 class="pt0">4.2 Is Rock Creek an Atmospheric Methane Source?</h3><p class="mb0">Rock Creek is a source of atmospheric methane, regardless of aeration status, or site. The air-water methane flux from Rock Creek varied between 0.2 and 1.5 mmol m<sup>−2</sup> d<sup>−1</sup> (note change in units to compare to literature values), which was similar to fluxes measured from several estuaries (<a href="#T3">Table 3</a>), and higher than those from oceanic environments, which vary between 0.0001 and 0.1 mmol m<sup>−2</sup> d<sup>−1</sup> (<a href="#B4">Bižić et al., 2020</a>). Rock Creek methane fluxes are on par with a shallow subarctic lake which reached almost 0.4 mmol m<sup>−2</sup> d<sup>−1</sup> (<a href="#B34">Jansen et al., 2020</a>), even though at times of ice-out, these lakes can release as much as 75 mmol m<sup>−2</sup> d<sup>−1</sup> (<a href="#B57">McIntosh Marcek et al., 2021</a>). Surface water concentrations were also similar to those measured from an aerated eutrophic lake (<a href="#B51">Martinez and Anderson, 2013</a>) suggesting methane is emitted in these aerated waters. A unique aspect to the work presented here is the high-frequency sampling over the warm season in Rock Creek which measured consistently high concentrations (400–1,000 nM) in the bottom water (<a href="#F11">Figure 11</a>) that rival what has been measured in the anoxic bottom waters of the mainstem Chesapeake Bay in mid-summer (<a href="#B24">Gelesh et al., 2016</a>) and further supports a sustained methane flux to the atmosphere. Thus this relatively shallow (∼3 m) eutrophic estuary, may contribute more methane than previously thought, as was the case for streams and rivers (<a href="#B76">Stanley et al., 2016</a>), and conforms to our understanding of coastal ecosystems as having an outsized influence on methane fluxes in a global context.</p><div class="DottedLine"></div><div class="Imageheaders">TABLE 3</div><div class="FigureDesc"><a href="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t003.jpg" name="Table3" target="_blank"> <picture> <source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=480&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t003.jpg" media="(max-width: 563px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=370&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t003.jpg" media="(max-width: 1024px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=290&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t003.jpg" media="(max-width: 1441px)"><source type="image/webp" srcset="https://images-provider.frontiersin.org/api/ipx/w=410&f=webp/https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t003.jpg" media=""><source type="image/jpg" srcset="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t003.jpg" media=""> <img src="https://www.frontiersin.org/files/Articles/866152/fenvs-10-866152-HTML-r1/image_m/fenvs-10-866152-t003.jpg" alt="www.frontiersin.org" id="T3" loading="lazy"> </picture> </a><p><strong>TABLE 3</strong>. Examples of estuarine flux of methane to the atmosphere.</p></div><div class="clear"></div><div class="DottedLine"></div><h3 class="pt0">4.3 The Source of Methane: All From Sediments?</h3><p class="mb0">Sedimentary methanogenesis is likely the main source of methane to the water column of Rock Creek because the highest dissolved methane concentrations were measured in the sediments, and methane concentrations in the bottom water were typically higher than the surface water. Biogenic methane is also supported with the sedimentary porewater methane δ<sup>13</sup>C-CH<sub>4</sub> values in the deep sediments being < −70‰ (<a href="#B81">Whiticar, 1999</a>). Yet, there was also evidence that methane produced in the deep sediments went through some degree of microbial oxidation before reaching the overlying water. The sediment porewater methane profiles showed classic concave-up shapes which are indicative of the anaerobic oxidation of methane (AOM) working in concert with sulfate reduction, as expressed here with the sulfate methane transition (SMT) depths (<a href="#B35">Jørgensen et al., 2019</a>). AOM is also supported with the porewater methane isotopic composition data. In 2019, the porewater δ<sup>13</sup>C-CH<sub>4</sub> values also increased up the core through the SMT depth. This pattern indicates AOM; as the methane diffuses along the concentration gradient, microbial communities preferentially utilize <sup>12</sup>C and leave the <sup>13</sup>C behind (thereby values increase) as methane is oxidized (<a href="#B81">Whiticar, 1999</a>). In 2019, this pattern is clear regardless of aeration status but there is a shift to more <sup>13</sup>C depleted values when aerators were OFF. This shift could represent enhanced microbial methane production when aerators were OFF but would need to be validated with other information such as diagenetic modeling (e.g., <a href="#B50">Martens et al., 1998</a>). The 2019 sedimentary profiles, measured under the low-turbulence diffuse system, support our classic understanding of biogeochemical zonation (<a href="#B23">Froelich et al., 1979</a>) and suggests that the sediments are diffusion dominated.</p><p class="mb0">The pattern in 2018 was not as clear, possibly due to sediment disturbance with vigorous aeration. The destratification system employed during 2018 involved a high-volume through flow of air that leads to substantial physical disturbance of the water-column and sediments. Such disturbance at the aerators translated into variable porewater methane concentration and the isotope patterns, compared to outside the aeration zone (<a href="#F5">Figure 5</a>). This implies that a simple, steady-state 1D, diffusion dominated interpretation of these profiles cannot be applied here because the destratification system could have driven substantial advective exchange between sediments and the water-column. It is interesting to note that at the aerators (RC1 and RC2), there was a depletion of <sup>13</sup>C up the core to the sediment-water interface. One way to inject such a depleted signature is by methanogenesis which could happen in this agitated system by bubbling out deep methane during aeration or methanogenesis in the surface layers using non-competitive substrates (<a href="#B2">Alperin et al., 1988</a>). More work would be needed to support or refute these possibilities.</p><p class="mb0">Methane from the sediments can either diffuse into the overlying water or bubble out <em>via</em> ebullition (methane oversaturated porewaters forming bubbles). The methane released from both of those processes would have different isotopic signatures. For example, for bubbles to form in the sediments and efflux, the methane concentration must be above saturation which occurs in the sediments deeper than ∼10 cmbsf, depending on the site and year (see <a href="#F5">Figures 5</a>, <a href="#F6">6</a>). At these depths, the δ<sup>13</sup>C-CH<sub>4</sub> values were between −80 and −85‰, so we would expect to see these values in the overlying water column (assuming a small amount of methane from the bubble equilibrates with the water). Bubbles captured in a shallow water column off North Carolina showed no isotopic fractionation when released to the water (<a href="#B13">Chanton and Martens, 1988</a>). However, if the methane is diffusing into the water from the sediment surface, that methane would carry a δ<sup>13</sup>C-CH<sub>4</sub> value similar to the surface sediment methane signature. For 2018, the δ<sup>13</sup>C-CH<sub>4</sub> values of water column methane were always higher than the surface sediment, regardless of site or aeration status. This suggests that there could be some methane oxidation at the very sediment surface that we are missing in our sediment measurements. This was the same situation when the aerators were OFF in 2019. Perhaps future work could focus on the sediment-water interface as an area of intense methane oxidation, be it either aerobic or anaerobic.</p><p class="mb0">The situation is a little different when the aerators were ON in 2019; the water column methane always had a lower δ<sup>13</sup>C-CH<sub>4</sub> value than the surficial sediments. This suggests that the methane was not simply diffusing in from the sediments, and that there must be another source of methane injecting isotopically depleted carbon; such as, sediment ebullition, methane being advected out of the sediments into the water column from the aerators, or microbial methane production in the water column. We don’t have data to support or refute the advective release of methane out of the sediments other than to say that in 2019, the aeration was much finer and thus most likely less advective than in 2018 so it seems unlikely. Plus, since we were not focused on determining the ebullitive flux, its hard to evaluate. However, in 2018, we had evidence of an ebullitive flux of methane from the sediments. First, the methane in the air above Rock Creek waters had an average δ<sup>13</sup>C-CH<sub>4</sub> value of −50‰ which is slightly depleted in <sup>13</sup>C from the global, well-mixed value from the northern hemisphere of −47.4‰ (<a href="#B40">Lan et al., 2021</a>). When this air δ<sup>13</sup>C-CH<sub>4</sub> value was compared to what was measured in the deep sediments, < −70‰, or in the water column, < −60‰, the <sup>13</sup>C depleted methane in air over Rock Creek waters can be explained with a small amount of biogenic methane from the sediments. Secondly, we measured a direct pulse of biogenic (δ<sup>13</sup>C-CH<sub>4</sub> = −80‰) methane in the air above RC7 at dawn when the aerators were OFF (<a href="#F7">Figure 7</a>). This was most likely due to ebullition from the sediments directly reaching the atmosphere which was trapped in the air above the water due to the air inversion that occurs at dawn (<a href="#B16">Crill et al., 1988</a>; <a href="#B60">Mukhophadhya et al., 2001</a>). While this observation of high concentrations of methane in the morning air is not unexpected, it clearly documents how aqueous environments contribute to the atmospheric methane isotopic signal which has recently been shown to be enhanced in biogenic methane sources (<a href="#B73">Schaefer et al., 2016</a>; <a href="#B40">Lan et al., 2021</a>). It also clearly shows that in order to fully capture methane dynamics in this system, future work should quantify the bubble flux from the tributary, much like was done in the temperature lake in California (<a href="#B51">Martinez and Anderson, 2013</a>).</p><h3 class="pt0">4.4 Evidence for Methane Production in the Water Column</h3><p class="mb0">While we might have captured methane bubbling out of Rock Creek in 2018, that process is very heterogeneous and not continuous, and may not fully explain the water column δ<sup>13</sup>C-CH<sub>4</sub> values from 2019 ON status. We thus explore the possibility of water column methane production under aeration. There is growing evidence for an oxic production pathway in surface waters (see review in <a href="#B4">Bižić et al., 2020</a>) that might be important. Most notably, it was concluded that 90% of methane in surface waters of a temperate lake was formed in the oxic surface waters, and not the sediments (<a href="#B21">Donis et al., 2017</a>). However, a recent reevaluation of this work concluded that a sedimentary methane flux from the sediments flanking the lake in the shallow waters could explain the oversaturated surface waters (<a href="#B64">Peeters et al., 2019</a>). In well stratified lakes, methane production in surface waters is shown to scale with sediment area and mixed layer volume (<a href="#B26">Günthel et al., 2019</a>). Lakes are hydrodynamically very different from estuaries, making study comparisons to Rock Creek tenuous. Furthermore, the hydrodynamic conditions in destratification systems, such as in 2018, can create large cells of overturning water that may impact a much larger area of the tidal system (<a href="#B25">Gibbs and Howard-Williams, 2018</a>). The lack of studies of sub-estuaries alone highlights the need to quantify the sources of methane from systems such as Rock Creek.</p><p class="mb0">The box model, which balances sedimentary and air-water methane fluxes, shows that for most of the time, regardless of if aerators were ON or OFF, there is an additional methane source in the water column in 2019 (<a href="#F11">Figure 11</a>). The observation that outside the aeration zone, RC7 and RC9, there is also an additional methane source could indicate the transport of methane from the Patapsco River, lateral inputs of water from across the creek as destratification cells pull in sources from a cross-section of the creek or simply upstream methane sources. Previous work in rivers also documented higher methane flux in upstream surface waters (<a href="#B18">de Angelis and Scranton, 1993</a>; <a href="#B1">Abril and Iversen, 2002</a>; <a href="#B58">Middelburg et al., 2002</a>); as well as higher methane in smaller width creeks (<a href="#B5">Borges and Abril, 2011</a>).</p><p class="mb0">The further observation that there is a methane source regardless of the status of the aerators (either ON or OFF) suggests there is a bigger ecosystem response than just the influx of oxygen from the aerators. One possibility could be the presence of algal blooms which produce methane as a byproduct (<a href="#B4">Bižić et al., 2020</a>). Such blooms have been shown to occur before when aerators are turned OFF (<a href="#B28">Harris et al., 2015</a>), and we also documented one right after the aerators were turned OFF in 2018 (<a href="#F10">Figure 10</a>). Increased surface water oxygen levels were observed with an increase in particulate organic nitrogen and carbon, and iron (data not shown). The OsmoSamplers also captured a pulse of methane right after this oxygen pulse. If these blooms are a typical phenomenon right after turning the aerators OFF, it could explain why the methane fluxes were higher 3 days after turning aerators OFF in 2021 (<a href="#F4">Figure 4D</a>) than when the aerators were ON. Future work could also directly quantify rates of methane production in aerobic waters.</p><h3 class="pt0">4.5 Evidence for Aerobic Methane Oxidation in the Water Column?</h3><p class="mb0">While it seems clear that the aerators are enhancing a methane flux to the air above the creek, that the sediments are the main source of methane to the waters and there possibly is methane production in these aerated waters, we also considered if there is any microbially mediated oxidation in the surface waters that might offset this atmospheric flux. The idea here is that the waters are being oxygenated due to the aeration, as the oxygen data suggested happened during the experiment (<a href="#F3">Figure 3</a>), and this oxygen is allowing for aerobic methane oxidation to happen. We have already speculated about a small potential oxidation process altering the δ<sup>13</sup>C-CH<sub>4</sub> from our surficial sedimentary methane values to the bottom water. Here, we focus solely on the water column. Since we didn’t directly measure microbial rates, we need to rely on geochemical data. We did this in four ways. First, we interrogated the stable carbon isotopes of methane measured in the water column. The δ<sup>13</sup>C-CH<sub>4</sub> values give a bulk measure of what has happened to that methane since it was formed; the bulk methane signature would be more enriched in <sup>13</sup>C if it had been oxidized. In order to do this, we considered that the methane in the bottom water is the source for the methane in the surface water. <a href="#F5">Figures 5</a>, <a href="#F6">6</a> water column isotope profiles clearly show that in all cases, the surface water is always slightly <sup>13</sup>C depleted, if at all different, contrary to what would be found if aerobic methane oxidation was occurring. Secondly, we looked for a correlation between oxygen concentrations and methane concentrations. For this correlation, we surmised that if oxygen was controlling methane levels, we would see a linear relationship between the two. There was no correlation (<a href="#SM1">Supplementary Figure S8A</a>), which suggests oxygen is not the limiting factor for aerobic methane oxidation. Thirdly, we considered that maybe aerobic methane oxidation was limited by dissolved inorganic nitrogen (DIN), as has been proposed in the literature (<a href="#B71">Sansone and Martens, 1978</a>). We surmise that if there is a negative relationship between DIN and water column methane concentrations, there could be evidence for aerobic methane oxidation. In other words, if methanotrophs were stimulated by DIN in an otherwise oxic water column, we would find low methane concentrations. This pattern did not emerge (<a href="#SM1">Supplementary Figure S8B</a>). And finally, in the box model approach described in Section 4.3, for most of the time in 2018 and 2019, there was a source of methane and not a sink. The exception was in 2019 when the aerators were ON. Taking all of this geochemical evidence into consideration, we conclude that we found little to no geochemical evidence for water column aerobic methane oxidation. Future work could focus on looking for microbial signature of oxidation in the water column.</p><h3 class="pt0">4.6 Complicating Factors With Experimental Design</h3><p class="mb0">We have discussed the differences in atmospheric methane flux between the 2018 and 2019 aeration experiments to most likely be a factor of the bubble size during aeration. The large bubble aerators resulted in a larger flux of methane to the atmosphere than the small bubble aerators. However there were also differences between the 2 years in how long the aerators were turned OFF. Despite variability in the concentrations, we found almost no effect of aeration when the aerators were turned OFF for a few days (2016, 2021), but there was modest methane accumulation as oxygen was depleted for 3–7 days in 2018 and significant methane increases after 13 days in 2019. Thus, it appears that methane accumulation requires more than a day to emerge after deoxygenation, despite prior studies in Arctic lake systems (<a href="#B57">McIntosh Marcek et al., 2021</a>) and in the Chesapeake Bay mainstem (<a href="#B24">Gelesh et al., 2016</a>) that appear to indicate that methane is immediately released from sediments following deoxygenation. However, the temporal resolution of prior measures likely cannot capture day to day dynamics, and there are few, if any real-time measures of methane and oxygen in coastal systems to confirm the time-scale of methane responses to deoxygenation.</p><p class="mb0">A second source of complication in our interpretations is that the experimental study is embedded in a system with background environmental variability. Thus, our “ON-OFF” study design using whole ecosystem manipulations does not offer a true controlled experimental system. For example, unlike previous work in Rock Creek (<a href="#B28">Harris et al., 2015</a>), the waters did not go anoxic within 1 day in 2018 and 2019 and only at RC2 in 2018 did anoxia ever emerge. In 2018, there was also an influx of oxygen-rich water that appeared in the aeration zone within days after the aerators were turned OFF and when oxygen should have been depleted (seen in time series data on <a href="#F10">Figure 10</a>). Although we did not measure methane during this event, the rapid increase in methane concentrations immediately after this pulse of tidally driven, oxygen-rich water could be the result of a rapid response to deoxygenation (that was not interrupted by the event) or an influx of methane rich water from upstream after the event. Our measurements in 2021 did indicate higher methane concentrations upstream of the aeration zone (∼1,000 nM; <a href="#F4">Figure 4D</a>), but these are substantially smaller than the concentrations measured after the event in 2018 (1,500 to 2,500 nM; <a href="#F10">Figure 10</a>), suggesting that an upstream source is unlikely. Other potential inputs of methane are currently unknown, such as groundwater, and there are no substantial tidal wetlands that could serve as a methane source. Although groundwater remains unlikely since chloride concentrations slightly increased with depth in sediment porewaters, our understanding of the drivers for the patterns remains elusive and long term observatories are needed.</p><p class="mb0">Ecosystem responses to aeration beyond deoxygenation also likely feedback to influence methane dynamics in the estuary. <a href="#B28">Harris et al. (2015)</a> reported a substantial algal bloom in Rock Creek within a week of the aerators being turned off, generating high surface water oxygen and organic matter concentrations. Given that Rock Creek has substantial light attenuation (Secchi Depth < 0.5), the vertical mixing induced by aerators likely keeps phytoplankton mixed below the photic layer, and when aerators are turned OFF, water-column stabilization allows for phytoplankton blooms. This is consistent with recent work in Chesapeake Bay that suggest that high chlorophyll-a packaging under low-light conditions combined with high nutrient concentrations allows for rapid algal growth when light becomes available (<a href="#B8">Buchanan, 2020</a>). Methane concentrations reached the highest levels we measured at RC2 in 2018 (∼2,700 nM), 3-days after an increase in particulate organic carbon of 700 μM after the aerators were turned OFF (data not shown). The consumption of this organic matter that likely followed may have generated substantial new methane, especially considering that oxygen concentrations were consistently below 32 μM (1 mg L<sup>−1</sup>) for much of the following week (<a href="#F10">Figure 10</a>). Given that the “destratification” approach used in 2018 was designed to physically mix and overturn the waters, such a phytoplankton response is likely. The “diffuse” approach used in 2019 involved much less physical disruption of the water-column (water-column salinity profiles were similar during ON and OFF), which may have prevented a phytoplankton response and also allowed for stable conditions that allowed the microbial communities to organize along expected redox conditions. This finding also points to an important impact of aerator design on our results, with differences in impact dependent on the mechanism of aeration and questions remain regarding how best to quantify mixing, oxygenation, and even water fluxes dependent on whether engineering impacted circulation (destratification design through large bubbles) versus diffusion of oxygen (small bubble aeration design).</p><a id="h6" name="h6"></a><h2>5 Conclusion</h2><p class="mb15">There is evidence that the sediments are the main source of methane in Rock Creek as conceptualized in <a href="#F1">Figure 1</a>, although we cannot rule out upstream creek waters or <em>in situ</em> production in the water column as additional sources. Our measurements suggest that aeration decreases the time frame available for aerobic methane oxidation in the water-column, thus connecting the sediment to the atmosphere more directly. This study did not allow for an in-depth examination of seasonal variations in methane sources and sinks, nor did it measure the impact of other complex factors on methane, such as shifting nutrient status and related microbial responses. Our experiments were also not intended to test aeration system design and operation (bubble size and density) in relation to methane flux, instead we used models to evaluate these impacts. Future work should 1) couple methane concentration and isotope data along with microbial rate measurements under various nutrient and oxygen conditions upstream and downstream of the aerators; 2) investigate seasonal changes to understand the complex factors controlling methane flux from this eutrophic estuary; and 3) characterize aerator design impacts on local hydrodynamics to better characterize physical effects on sediment transport and sediment-water exchange.</p><p class="mb15">Keeping these recommendations in mind, the strength of our study can be distilled into two central findings we emphasize here: 1) The shallow, Rock Creek sub-estuary is a source of methane to the atmosphere, regardless of engineering intervention, and methane production and flux is likely enhanced as a consequence of eutrophication and 2) the strength of the methane flux is impacted by aeration bubble size design. Our conceptual model lays out the processes and impacts that are connected to these findings (<a href="#F1">Figure 1</a>). The smallest fluxes occurred with the small bubble system (<a href="#F9">Figure 9</a>), and we predict that continued implementation of this system combined with potential future oligotrophication could reduce methane fluxes. The largest fluxes occurred with the older, large bubble system that was intended to encourage destratification. The data support the hypothesis that aeration can lead to higher atmospheric methane fluxes and that aerator design is crucial to mitigating methane transfer. A key motivation for this study was to investigate the potential for unintended consequences of this engineering intervention in relation to greenhouse gas emissions. The dependency of the measured methane atmospheric fluxes on bubble size suggests that there is a path forward towards optimizing aerator design to reduce this consequence in eutrophic tidal waters (e.g., implement small bubble aeration). This study also adds to a growing body of literature quantifying methane fluxes in coastal waters. Regardless of aerator status or design, the current condition of Rock Creek as a eutrophic ecosystem characterized by high primary production of organic matter impacts its overall role as a source of methane to the atmosphere. As coastal water quality policies are implemented and managers seek both solutions and greater understanding of the complex biogeochemistry that impacts restoration trajectories in eutrophic systems, work on both engineering solutions and interpretation of restoration monitoring data would benefit from including methane dynamics and greenhouse gas impacts into holistic management frameworks.</p><a id="h7" name="h7"></a><h2>Data Availability Statement</h2><p class="mb15">The datasets presented in this study can be found in online repositories. The names of the repository/repositories and accession number(s) can be found below: <a href="https://www.ncei.noaa.gov/archive/accession/0244510">https://www.ncei.noaa.gov/archive/accession/0244510</a>.</p><a id="h8" name="h8"></a><h2>Author Contributions</h2><p class="mb0">LL, AH, JT, and LH conceived the idea; LL, CH, MF, CS, and EH conducted the experiment and analyzed samples; LL wrote the initial manuscript, and all authors contributed significant edits to the manuscript. We thank Mark Mobley for help with the bubble model, and the reviewers and editors who greatly improved this manuscript.</p><a id="h9" name="h9"></a><h2>Funding</h2><p class="mb0">Financial support for this project was through U.S. National Science Foundation CBET-1706416 (LH, AH, JT, and LL), Maryland Sea Grant (NA14OAR4170090, SA75281450-O; LH and JT), and Maryland SeaGrant REU program (NSF grant OCE-1262374).</p><a id="h10" name="h10"></a><h2>Conflict of Interest</h2><p class="mb0">The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.</p><a id="h11" name="h11"></a><h2>Publisher’s Note</h2><p class="mb15">All claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. Any product that may be evaluated in this article, or claim that may be made by its manufacturer, is not guaranteed or endorsed by the publisher.</p><a id="h12" name="h12"></a><h2>Acknowledgments</h2><p class="mb03">We thank Anne Arundel county for access to the aerators, the residents of Rock Creek for allowing us to conduct the experiments in their backyard, and Janis Markusic for her support. The work could not have been done without the field help of Zachary Gotthardt, Lillian Henderson, and Brittany Clark and laboratory help of Maureen Strauss and Hadley McIntosh Marcek. We thank Dave Oliff at the Florida State University machine shop for building the fast OsmoSamplers. Work on this paper by Lapham was possible with the expertise and care of healthcare workers at the Johns Hopkins Breast Cancer Center and support from Marcia Lapham and Dr. Cédric Magen. Co-author Harris was supported by the Rock Creek team as a nursing mother in the field with patience, privacy, and good humor during boat work with co-authors Szewczyk and Testa. Lapham, Harris, and Testa depended on childcare provided by their communities and especially their spouses during the COVID-19 pandemic and are thankful to have had the time for this research. This is UMCES contribution #6188.</p><a id="h13" name="h13"></a><h2>Supplementary Material</h2><p class="mb15" id="SM1">The Supplementary Material for this article can be found online at: <a href="https://www.frontiersin.org/articles/10.3389/fenvs.2022.866152/full#supplementary-material">https://www.frontiersin.org/articles/10.3389/fenvs.2022.866152/full#supplementary-material</a></p><a id="h14" name="h14"></a><h2>References</h2><div class="References"><p class="ReferencesCopy1"><a name="B1" id="B1"></a>Abril, G., and Iversen, N. (2002). Methane Dynamics in a Shallow Non-tidal Estuary (Randers Fjord, Denmark). <em>Mar. 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Data</em> 21 (1), 78–80. doi:10.1021/je60068a029</p><p class="ReferencesCopy2"><a href="https://doi.org/10.1021/je60068a029">CrossRef Full Text</a> | <a href="https://scholar.google.com/scholar?hl=en&as_sdt=0%2C5&q=Solubility+of+Methane+in+Distilled+Water+and+Seawater&btnG=">Google Scholar</a></p></div></div><div class="thinLineM20"></div><div class="AbstractSummary"><p><span>Keywords:</span> methane, aeration, eutrophication, estuary, OsmoSampler</p><p><span>Citation:</span> Lapham LL, Hobbs EA, Testa JM, Heyes A, Forsyth MK, Hodgkins C, Szewczyk C and Harris LA (2022) The Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary. <em>Front. Environ. Sci.</em> 10:866152. doi: 10.3389/fenvs.2022.866152</p><p id="timestamps"><span>Received:</span> 30 January 2022; <span>Accepted:</span> 13 June 2022;<br><span>Published:</span> 11 August 2022.</p><div><p>Edited by:</p> <a href="https://loop.frontiersin.org/people/606160/overview">Daniel F. McGinnis</a>, Université de Genève, Switzerland</div><div><p>Reviewed by:</p> <a href="https://loop.frontiersin.org/people/1668703/overview">Peter Tango</a>, United States Geological Survey (USGS), United States<br><a href="https://loop.frontiersin.org/people/218446/overview">Wei-Dong Zhai</a>, Shandong University, China</div><p><span>Copyright</span> © 2022 Lapham, Hobbs, Testa, Heyes, Forsyth, Hodgkins, Szewczyk and Harris. This is an open-access article distributed under the terms of the <a rel="license" href="http://creativecommons.org/licenses/by/4.0/" target="_blank">Creative Commons Attribution License (CC BY).</a> The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.</p><p><span>*Correspondence:</span> Laura L. Lapham, <a href="mailto:lapham@umces.edu">lapham@umces.edu</a></p><div class="clear"></div></div></div></div> <p class="AbstractSummary__disclaimer"><span>Disclaimer: </span> All claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. 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General Commentary articles as well as Book Reviews in Frontiers in Environmental Science are only accepted upon invitation.\u003C\u002Fp\u003E\n\n\u003Cp\u003EFrontiers in Environmental Science is committed to advancing developments in the field through unrestricted access to articles, and assisting in the formulation of policies, by offering sound scientific evidence on environmental science, that will lead to a more sustainable world for the generations to come.\u003C\u002Fp\u003E",palette:"green",impactFactor:"4.6",citeScore:"3.1",citations:"54700",showTagline:e,twitter:"@FrontEnvSci",__typename:"Journal"},currentFrontiersJournal:{id:v,name:s,slug:w,printISSN:e,shortName:H,electronicISSN:I,abbreviation:U,specialtyId:e,publicationDate:e,isOnline:h,isOpenForSubmissions:h,spaceId:c,field:{id:V,domainId:c,__typename:W},__typename:a},articleHubSlug:f,articleHubPage:J,currentArticle:{id:866152,doi:X,title:Y,acceptanceDate:new Date(1655135875000),receptionDate:new Date(1643579715000),publicationDate:new Date(1660176000000),isPublished:h,abstract:Z,researchTopic:{id:22659,title:"Physical and Biogeochemical Processes Driving Methane Sources, Sinks and Emissions in Aquatic Systems: The Past, Present and Future under Global Change",articlesCount:_,isMagazinePage:k,slug:"physical-and-biogeochemical-processes-driving-methane-sources-sinks-and-emissions-in-aquatic-systems-the-past-present-and-future-under-global-change",isOpenForSubmission:k},articleType:{id:_,name:"Original Research"},stage:{id:K,name:f},keywords:["Methane","Aeration","Eutrophication","estuary","OsmoSampler"],authors:[{id:$,firstName:aa,lastName:"Lapham",givenNames:aa,isCorresponding:k,isProfilePublic:h,userId:$,affiliations:[{organizationName:o,countryName:n,cityName:f,stateName:f,zipCode:f}]},{id:ab,firstName:ac,lastName:"Hobbs",givenNames:ac,isCorresponding:k,isProfilePublic:h,userId:ab,affiliations:[{organizationName:o,countryName:n,cityName:f,stateName:f,zipCode:f}]},{id:ad,firstName:ae,lastName:"Testa",givenNames:ae,isCorresponding:k,isProfilePublic:h,userId:ad,affiliations:[{organizationName:o,countryName:n,cityName:f,stateName:f,zipCode:f}]},{id:af,firstName:ag,lastName:"Heyes",givenNames:ag,isCorresponding:k,isProfilePublic:h,userId:af,affiliations:[{organizationName:o,countryName:n,cityName:f,stateName:f,zipCode:f}]},{id:m,firstName:ah,lastName:"Forsyth",givenNames:ah,isCorresponding:k,isProfilePublic:k,userId:m,affiliations:[{organizationName:o,countryName:n,cityName:f,stateName:f,zipCode:f}]},{id:m,firstName:ai,lastName:"Hodgkins",givenNames:ai,isCorresponding:k,isProfilePublic:k,userId:m,affiliations:[{organizationName:o,countryName:n,cityName:f,stateName:f,zipCode:f}]},{id:aj,firstName:ak,lastName:"Szewczyk",givenNames:ak,isCorresponding:k,isProfilePublic:h,userId:aj,affiliations:[{organizationName:o,countryName:n,cityName:f,stateName:f,zipCode:f}]},{id:al,firstName:am,lastName:"Harris",givenNames:am,isCorresponding:k,isProfilePublic:h,userId:al,affiliations:[{organizationName:o,countryName:n,cityName:f,stateName:f,zipCode:f}]}],editors:[{id:an,firstName:"Daniel",lastName:"McGinnis",givenNames:"Daniel F.",isCorresponding:k,isProfilePublic:h,userId:an,affiliations:[{organizationName:"University of Geneva",countryName:"Switzerland",cityName:f,stateName:f,zipCode:f}]}],reviewers:[{id:ao,firstName:ap,lastName:"Zhai",givenNames:ap,isCorresponding:k,isProfilePublic:h,userId:ao,affiliations:[{organizationName:"Southern Marine Science and Engineering Guangdong Laboratory (Zhuhai)",countryName:"China",cityName:f,stateName:f,zipCode:f}]},{id:aq,firstName:ar,lastName:"Tango",givenNames:ar,isCorresponding:k,isProfilePublic:h,userId:aq,affiliations:[{organizationName:"United States Geological Survey (USGS), United States Department of the Interior",countryName:n,cityName:f,stateName:f,zipCode:f}]}],journal:{id:v,slug:w,name:s,shortName:H,electronicISSN:I,field:{id:V,domainId:c,__typename:W},specialtyId:e,journalSectionPaths:[{section:as,__typename:"journal_journalSectionPath"}],__typename:a},section:as,impactMetrics:{views:1296,downloads:503,citations:c},volume:aw,articleVolume:"Volume 10 - 2022",relatedArticles:[],isPublishedV2:k,contents:{fullTextHtml:"\u003Cdiv class=\"JournalAbstract\"\u003E\u003Ca id=\"h1\" name=\"h1\"\u003E\u003C\u002Fa\u003E\u003Ch1\u003EThe Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary\u003C\u002Fh1\u003E\u003Cdiv class=\"authors\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Fpeople\u002Fu\u002F23188\" class=\"user-id-23188\"\u003E\u003Cimg class=\"pr5\" src=\"https:\u002F\u002Floop.frontiersin.org\u002Fimages\u002Fprofile\u002F23188\u002F24\" onerror=\"this.src='http:\u002F\u002F3b76aaf63d1816bb57bf-a34624e694c43cdf8b40aa048a644ca4.r96.cf2.rackcdn.com\u002FDesign\u002FImages\u002Fnewprofile_default_profileimage_new.jpg'\" alt=\"www.frontiersin.org\"\u002F\u003ELaura L. Lapham\u003C\u002Fa\u003E*, \u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Fpeople\u002Fu\u002F1413210\" class=\"user-id-1413210\"\u003E\u003Cimg class=\"pr5\" src=\"https:\u002F\u002Floop.frontiersin.org\u002Fimages\u002Fprofile\u002F1413210\u002F24\" onerror=\"this.src='http:\u002F\u002F3b76aaf63d1816bb57bf-a34624e694c43cdf8b40aa048a644ca4.r96.cf2.rackcdn.com\u002FDesign\u002FImages\u002Fnewprofile_default_profileimage_new.jpg'\" alt=\"www.frontiersin.org\"\u002F\u003EEdward A. Hobbs\u003C\u002Fa\u003E, \u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Fpeople\u002Fu\u002F435564\" class=\"user-id-435564\"\u003E\u003Cimg class=\"pr5\" src=\"https:\u002F\u002Floop.frontiersin.org\u002Fimages\u002Fprofile\u002F435564\u002F24\" onerror=\"this.src='http:\u002F\u002F3b76aaf63d1816bb57bf-a34624e694c43cdf8b40aa048a644ca4.r96.cf2.rackcdn.com\u002FDesign\u002FImages\u002Fnewprofile_default_profileimage_new.jpg'\" alt=\"www.frontiersin.org\"\u002F\u003EJeremy M. Testa\u003C\u002Fa\u003E, \u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Fpeople\u002Fu\u002F1675968\" class=\"user-id-1675968\"\u003E\u003Cimg class=\"pr5\" src=\"https:\u002F\u002Floop.frontiersin.org\u002Fimages\u002Fprofile\u002F1675968\u002F24\" onerror=\"this.src='http:\u002F\u002F3b76aaf63d1816bb57bf-a34624e694c43cdf8b40aa048a644ca4.r96.cf2.rackcdn.com\u002FDesign\u002FImages\u002Fnewprofile_default_profileimage_new.jpg'\" alt=\"www.frontiersin.org\"\u002F\u003EAndrew Heyes\u003C\u002Fa\u003E, \u003Cimg class=\"pr5\" src=\"http:\u002F\u002F3b76aaf63d1816bb57bf-a34624e694c43cdf8b40aa048a644ca4.r96.cf2.rackcdn.com\u002FDesign\u002FImages\u002Fnewprofile_default_profileimage_new.jpg\" alt=\"www.frontiersin.org\"\u002F\u003EMelinda K. Forsyth, \u003Cimg class=\"pr5\" src=\"http:\u002F\u002F3b76aaf63d1816bb57bf-a34624e694c43cdf8b40aa048a644ca4.r96.cf2.rackcdn.com\u002FDesign\u002FImages\u002Fnewprofile_default_profileimage_new.jpg\" alt=\"www.frontiersin.org\"\u002F\u003ECasey Hodgkins, \u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Fpeople\u002Fu\u002F1710495\" class=\"user-id-1710495\"\u003E\u003Cimg class=\"pr5\" src=\"https:\u002F\u002Floop.frontiersin.org\u002Fimages\u002Fprofile\u002F1710495\u002F24\" onerror=\"this.src='http:\u002F\u002F3b76aaf63d1816bb57bf-a34624e694c43cdf8b40aa048a644ca4.r96.cf2.rackcdn.com\u002FDesign\u002FImages\u002Fnewprofile_default_profileimage_new.jpg'\" alt=\"www.frontiersin.org\"\u002F\u003ECurtis Szewczyk\u003C\u002Fa\u003E and \u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Fpeople\u002Fu\u002F1264726\" class=\"user-id-1264726\"\u003E\u003Cimg class=\"pr5\" src=\"https:\u002F\u002Floop.frontiersin.org\u002Fimages\u002Fprofile\u002F1264726\u002F24\" onerror=\"this.src='http:\u002F\u002F3b76aaf63d1816bb57bf-a34624e694c43cdf8b40aa048a644ca4.r96.cf2.rackcdn.com\u002FDesign\u002FImages\u002Fnewprofile_default_profileimage_new.jpg'\" alt=\"www.frontiersin.org\"\u002F\u003ELora A. Harris\u003C\u002Fa\u003E\u003C\u002Fdiv\u003E\u003Cul class=\"notes\"\u003E\u003Cli\u003EChesapeake Biological Laboratory, University of Maryland Center for Environmental Science, Solomons, MD, United States\u003C\u002Fli\u003E\u003C\u002Ful\u003E\u003Cp class=\"mb15\"\u003EEngineered aeration is one solution for increasing oxygen concentrations in highly eutrophic estuaries that undergo seasonal hypoxia. Although there are various designs for engineered aeration, all approaches involve either destratification of the water column or direct injection of oxygen or air through fine bubble diffusion. To date, the effect of either approach on estuarine methane dynamics remains unknown. Here we tested the hypotheses that 1) bubble aeration will strip the water of methane and enhance the air-water methane flux to the atmosphere and 2) the addition of oxygen to the water column will enhance aerobic methane oxidation in the water column and potentially offset the air-water methane flux. These hypotheses were tested in Rock Creek, Maryland, a shallow-water sub-estuary to the Chesapeake Bay, using controlled, ecosystem-scale deoxygenation experiments where the water column and sediments were sampled in mid-summer, when aerators were ON, and then 1, 3, 7, and 13 days after the aerators were turned OFF. Experiments were performed under two system designs, large bubble and fine bubble approaches, using the same observational approach that combined discrete water sampling, long term water samplers (OsmoSamplers) and sediment porewater profiles. Regardless of aeration status, methane concentrations reached as high as 1,500 nmol L\u003Csup\u003E−1\u003C\u002Fsup\u003E in the water column during the experiments and remained near 1,000 nmol L\u003Csup\u003E−1\u003C\u002Fsup\u003E through the summer and into the fall. Since these concentrations are above atmospheric equilibrium of 3 nmol L\u003Csup\u003E−1\u003C\u002Fsup\u003E, these data establish the sub-estuary as a source of methane to the atmosphere, with a maximum atmospheric flux as high as 1,500 µmol m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E, which is comparable to fluxes estimated for other estuaries. Air-water methane fluxes were higher when the aerators were ON, over short time frames, supporting the hypothesis that aeration enhanced the atmospheric methane flux. The fine-bubble approach showed lower air-water methane fluxes compared to the larger bubble, destratification system. We found that the primary source of the methane was the sediments, however, \u003Cem\u003Ein situ\u003C\u002Fem\u003E methane production or an upstream methane source could not be ruled out. Overall, our measurements of methane concentrations were consistently high in all times and locations, supporting consistent methane flux to the atmosphere.\u003C\u002Fp\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"JournalFullText\"\u003E\u003Ca id=\"h2\" name=\"h2\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003E1 Introduction\u003C\u002Fh2\u003E\u003Cp class=\"mb15\"\u003EThe eutrophication of estuaries as a result of nutrient enrichment is a global phenomenon, with consequences that include deoxygenation and hypoxia (\u003Ca href=\"#B19\"\u003EDiaz and Rosenberg, 2008\u003C\u002Fa\u003E). In fact, modeling and data analysis suggests that dissolved oxygen in estuaries will continue to decline into the future, primarily as a result of long-term warming (\u003Ca href=\"#B7\"\u003EBreitburg et al., 2018\u003C\u002Fa\u003E; \u003Ca href=\"#B62\"\u003ENi et al., 2019\u003C\u002Fa\u003E; \u003Ca href=\"#B82\"\u003EWhitney and Vlahos, 2021\u003C\u002Fa\u003E). The primary mitigation tool has been to enforce managed reductions of land-based nutrients in the United States (\u003Ca href=\"#B46\"\u003ELinker et al., 2013\u003C\u002Fa\u003E) and in Europe (\u003Ca href=\"#B30\"\u003EHELCOM, 2021\u003C\u002Fa\u003E), yet engineered solutions are also being considered (\u003Ca href=\"#B14\"\u003EConley et al., 2009\u003C\u002Fa\u003E; \u003Ca href=\"#B44\"\u003ELehtoranta et al., 2022\u003C\u002Fa\u003E). Engineered aeration efforts work by either destratifying the water column or directly injecting oxygen to the water (\u003Ca href=\"#B28\"\u003EHarris et al., 2015\u003C\u002Fa\u003E; \u003Ca href=\"#B78\"\u003EStigebrandt et al., 2015\u003C\u002Fa\u003E; \u003Ca href=\"#B39\"\u003EKoweek et al., 2020\u003C\u002Fa\u003E). This has also been commonly done in small lake systems (e.g., \u003Ca href=\"#B51\"\u003EMartinez and Anderson, 2013\u003C\u002Fa\u003E; \u003Ca href=\"#B32\"\u003EHounshell et al., 2021\u003C\u002Fa\u003E) and reservoirs (\u003Ca href=\"#B55\"\u003EMcCord et al., 2016\u003C\u002Fa\u003E). While aeration should relieve the low oxygen problem to create habitat for metazoan life, prevent the noxious release of sulfide from sediments, and enhance coupled nitrification-denitrification, an additional potential consequence is that aeration could also enhance atmospheric methane emissions in estuaries. If this is true, methane emissions from estuaries that undergo aeration could be larger than currently considered in the global budget (\u003Ca href=\"#B72\"\u003ESaunois et al., 2020\u003C\u002Fa\u003E). It is critical to constrain all sources of methane to the atmosphere since it is a powerful greenhouse gas (\u003Ca href=\"#B22\"\u003EForster et al., 2007\u003C\u002Fa\u003E; \u003Ca href=\"#B20\"\u003EDlugokencky, 2020\u003C\u002Fa\u003E). However, studies from an aerated freshwater reservoir show that the methane emissions were lower than a nearby natural reservoir (\u003Ca href=\"#B53\"\u003EMcClure et al., 2018\u003C\u002Fa\u003E; \u003Ca href=\"#B54\"\u003EMcClure et al., 2021\u003C\u002Fa\u003E). To date, this interplay between engineered aeration and methane fluxes in a natural estuary has not been rigorously tested.\u003C\u002Fp\u003E\u003Cp class=\"mb15\"\u003EEstuaries are dynamic environments, generating temporally and spatially-varying habitats in which methane producing and consuming processes occur. The primary source of methane is the underlying sediments, as in most organic rich environments (\u003Ca href=\"#B49\"\u003EMartens and Berner, 1974\u003C\u002Fa\u003E; \u003Ca href=\"#B67\"\u003EReeburgh, 2007\u003C\u002Fa\u003E). However, there is an increasing appreciation for alternative sources such as demethylation of organic phosphonates (\u003Ca href=\"#B37\"\u003EKarl et al., 2008\u003C\u002Fa\u003E), bacterial degradation of water column dissolved organic matter (\u003Ca href=\"#B68\"\u003ERepeta et al., 2016\u003C\u002Fa\u003E) and\u002For production by phytoplankton (\u003Ca href=\"#B4\"\u003EBižić et al., 2020\u003C\u002Fa\u003E) that have not been fully explored in estuarine systems. In shallow-water, dynamic coastal and estuarine environments, methane can also be delivered with currents or tides from lateral sources (\u003Ca href=\"#B4\"\u003EBižić et al., 2020\u003C\u002Fa\u003E). While methane formed in the sediments can enter the water column through ebullition (\u003Ca href=\"#B6\"\u003EBoudreau, 2012\u003C\u002Fa\u003E) or diffusion, nearly 85% of the methane produced within sediments is oxidized anaerobically before it reaches the sediment-water interface \u003Cem\u003Evia\u003C\u002Fem\u003E microbially mediated reactions including sulfate reduction, nitrate reduction, and iron reduction (\u003Ca href=\"#B23\"\u003EFroelich et al., 1979\u003C\u002Fa\u003E; \u003Ca href=\"#B67\"\u003EReeburgh, 2007\u003C\u002Fa\u003E). The remaining methane released from the sediments to the overlying water column can then be oxidized aerobically \u003Cem\u003Evia\u003C\u002Fem\u003E methanotrophs (\u003Ca href=\"#B27\"\u003EHanson and Hanson, 1996\u003C\u002Fa\u003E). Thus a conceptual model for a healthy estuary shows a small methane flux to the atmosphere (\u003Ca href=\"#F1\"\u003EFigure 1A\u003C\u002Fa\u003E). Alternatively, when estuarine waters are highly eutrophic, there is a breakdown in the aerobic biofilter in the water column and this results in an enhanced methane flux when the bottom waters go hypoxic and anoxic (\u003Ca href=\"#B24\"\u003EGelesh et al., 2016\u003C\u002Fa\u003E). Thus, under these conditions, there is a higher methane flux to the atmosphere (\u003Ca href=\"#F1\"\u003EFigure 1B\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 1\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g001.jpg\" name=\"Figure1\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g001.gif\" id=\"F1\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 1\u003C\u002Fstrong\u003E. Conceptual diagrams of methane dynamics in estuarine waters. With no aeration, \u003Cstrong\u003E(A)\u003C\u002Fstrong\u003E the lowest methane flux comes from a water column that is well oxygenated and \u003Cstrong\u003E(B)\u003C\u002Fstrong\u003E there is a moderate flux when waters are eutrophic. When aerated, the methane flux is \u003Cstrong\u003E(C)\u003C\u002Fstrong\u003E highest when there are large bubbles and \u003Cstrong\u003E(D)\u003C\u002Fstrong\u003E moderate when there are small bubbles.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"mb15\"\u003EUnder this simple conceptual model, it is enticing to speculate that if the bottom waters were re-oxygenated, this would return the aerobic biofilter to its normal state and lower the methane flux to the atmosphere. However, methane is a highly insoluble gas, and the mere addition of air bubbles (devoid of methane) and physical movement of the water during aeration could instead promote methane to dissolve into the rising bubbles and released to the atmosphere. In this case, the size of the air bubbles injected into the bottom waters could affect the magnitude of the atmospheric flux. For example, when the bubbles are large, the physical movement of the turbulent water would release large amounts of methane (\u003Ca href=\"#F1\"\u003EFigure 1C\u003C\u002Fa\u003E). When the bubbles are small, there is the possibility for some of the oxygen to diffuse into the surrounding water (which is the goal of aeration systems) and promote aerobic methane oxidation (\u003Ca href=\"#F1\"\u003EFigure 1D\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cp class=\"mb15\"\u003EOxygenated water columns are necessary for aerobic methane oxidizing bacteria to help control methane emissions to the atmosphere. In the simplest case of the open ocean where there are deep, well-oxygenated waters, nearly all the dissolved methane in the water column is oxidized (\u003Ca href=\"#B45\"\u003ELeonte et al., 2017\u003C\u002Fa\u003E; \u003Ca href=\"#B65\"\u003EPohlman et al., 2017\u003C\u002Fa\u003E). Yet even in these systems, studies have shown there is a lag time for aerobic oxidation (\u003Ca href=\"#B12\"\u003EChan et al., 2019\u003C\u002Fa\u003E), with notable exceptions such as the rapid response of methanotrophs to methane released during events like the Deepwater Horizon oil spill (\u003Ca href=\"#B38\"\u003EKessler et al., 2011\u003C\u002Fa\u003E). Classic works have shown the highest rates of aerobic methane oxidation at the oxycline in arctic lakes (\u003Ca href=\"#B70\"\u003ERudd and Hamilton, 1978\u003C\u002Fa\u003E) suggesting that these organisms are facultative microaerophiles who work in these strong oxygen gradients (\u003Ca href=\"#B63\"\u003EOswald et al., 2015\u003C\u002Fa\u003E; \u003Ca href=\"#B77\"\u003ESteinle et al., 2017\u003C\u002Fa\u003E). Aerobic methane oxidation can also proceed at high oxygen concentrations, especially when nitrogen is available. Under eutrophic conditions, when dissolved inorganic nitrogen concentrations >20 µM (M is used throughout as a symbol for mol L\u003Csup\u003E−1\u003C\u002Fsup\u003E), aerobic methane oxidation occurs at oxygen levels >31 µM; much higher than found in the oxyclines (\u003Ca href=\"#B71\"\u003ESansone and Martens, 1978\u003C\u002Fa\u003E). Along with nitrogen, micronutrients (especially copper) stimulate activity of methane oxidizing bacteria (\u003Ca href=\"#B75\"\u003ESemrau et al., 2010\u003C\u002Fa\u003E). Not surprisingly the dissolved methane concentration in the environment has been shown to influence the rate of water column methane oxidation. For example, in Arctic waters, higher rates of methane oxidation were measured when higher concentrations of methane were available in summer (\u003Ca href=\"#B52\"\u003EMau et al., 2013\u003C\u002Fa\u003E). However, during the Deepwater Horizon oil spill, aerobic methane oxidation rates decreased over time after an initial spike, even though methane concentrations remained high (\u003Ca href=\"#B15\"\u003ECrespo-Medina et al., 2014\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cp class=\"mb15\"\u003EHere we present a study that quantified the effect of engineered aeration on air-water methane emissions from a eutrophic estuary. We leveraged a unique opportunity to manipulate whole ecosystem dissolved oxygen concentrations using two types of engineered destratification systems, that we distinguish based on the bubble size (large bubble and small bubble aeration). These systems provided an ideal opportunity to test the hypothesis that the physical disturbance introduced to a eutrophic system with bubbles will enhance the atmospheric methane flux (\u003Ca href=\"#F1\"\u003EFigures 1C,D\u003C\u002Fa\u003E), regardless of oxygen concentration. We also surmised that the two different destratification systems would result in different methane fluxes to the atmosphere based on their bubble size; where the small bubbles would dissolve before reaching the air-water interface and not act as a transfer mechanism like larger bubbles. Furthermore, since the very idea behind engineered aeration is to add oxygen to the water column, we also hypothesized that aeration could enhance aerobic methane oxidation in the water column, which could act to lower the flux of methane to the atmosphere. To test our hypotheses, we conducted experimental manipulations of the aeration systems and sampled surface and bottom waters with aerators ON and then 1–13 days after the aerators were turned OFF. We hypothesized the source of methane in the water was the sediments, thus we also collected sediment cores at the same time. To put the discrete, experimental time points into a longer term context, we also present unique time-series measurements of methane concentrations in bottom waters across the whole estuary. Ultimately, this study contributes data to the growing literature of methane dynamics in shallow, eutrophic environments.\u003C\u002Fp\u003E\u003Ca id=\"h3\" name=\"h3\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003E2 Materials and Methods\u003C\u002Fh2\u003E\u003Ch3 class=\"pt0\"\u003E2.1 Field Description and Experimental Setup\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EThe experiments were carried out in a 353-ha tidal tributary to the Chesapeake Bay found in Anne Arundel County, Maryland (\u003Ca href=\"#F2\"\u003EFigure 2\u003C\u002Fa\u003E). Rock Creek’s watershed is 80% residential and 20% forested. Due to the poor water quality over the past few decades (e.g., anoxia and extensive algal blooms), the county installed an engineered destratification system in 1988 to bring dissolved oxygen back to former levels (\u003Ca href=\"#B28\"\u003EHarris et al., 2015\u003C\u002Fa\u003E). Herein, we call this the “large bubble aeration” system. Up until 2019, this system was made up of 138 ultra-coarse air diffusers distributed along a pipe that lines the middle creek channel with ∼20 mm bubble size (\u003Ca href=\"#B11\"\u003ECH2M_Hill, 2011\u003C\u002Fa\u003E), as described in design specifications reported by \u003Ca href=\"#B17\"\u003EDames and Moore (1988)\u003C\u002Fa\u003E. The goal of the system was to vigorously overturn the water column in order to continuously introduce oxygen \u003Cem\u003Evia\u003C\u002Fem\u003E de-stratification into the bottom waters, and the system was run continuously throughout the day. Every year, the aerators are turned on June 1 and turned off October 1, in order to minimize the effects of summertime hypoxia. An early study of this system determined that the zone of aeration influence was ∼74 ha and that bottom waters remained oxic when diffusers were ON, but became anoxic within one tidal cycle when aerators were OFF (\u003Ca href=\"#B28\"\u003EHarris et al., 2015\u003C\u002Fa\u003E; \u003Ca href=\"#B29\"\u003EHarris et al., 2016\u003C\u002Fa\u003E). In spring 2019, the aeration system was upgraded to fine bubble diffusers to provide more oxygen to the water column, herein referred to as the “small bubble aeration” system. The goal of the diffusers is primarily to overturn the water to allow re-aeration to occur at the water surface, not to add oxygen from the bubbles themselves. There are two 213 m long diffusers emanating from the shore mainline, which provide air at a rate of 180 scfm (standard cubic feet per minute) in a continuous bubble pattern (0.26 scfm ft\u003Csup\u003E−1\u003C\u002Fsup\u003E) of bubbles 3 mm in diameter. The surface water expression of the new aeration system is shown in \u003Ca href=\"#F2\"\u003EFigure 2B\u003C\u002Fa\u003E. With this new system, the county public works turns OFF the aerators every night to reduce neighborhood noise and energy consumption.\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 2\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g002.jpg\" name=\"Figure2\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g002.gif\" id=\"F2\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 2\u003C\u002Fstrong\u003E. Sample location map for Rock Creek. Red lines show location of the aeration tubes, black dots are water column and sediment station locations, white stars are OsmoSampler locations (very close to black dots), and black cross is location of dock where the benthic lander was deployed. \u003Cstrong\u003E(A)\u003C\u002Fstrong\u003E Location of Rock Creek (black square) in the northern Chesapeake Bay near Baltimore, Maryland. \u003Cstrong\u003E(B)\u003C\u002Fstrong\u003E Photo of bubbles breaking the surface in 2019 (photo by Laura Lapham).\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"mb0\"\u003EOver the course of 4 years (2016, 2018, 2019, 2021), the water column and sediments were sampled along the creek both within and outside the aeration zone and with both engineering designs (\u003Ca href=\"#T1\"\u003ETables 1\u003C\u002Fa\u003E, \u003Ca href=\"#T2\"\u003E2\u003C\u002Fa\u003E). Stations within the aeration zone included RC1, which is located at the up-creek limit of aeration, and RC2 which is found mid-channel and directly in the aeration zone (\u003Ca href=\"#F2\"\u003EFigure 2\u003C\u002Fa\u003E). RC7 is ∼1 km downstream from the end of aeration, still within the zone of influence of aeration and where the creek widens, and RC9b is a background site, close to the mouth of the Patapsco River and outside the zone of influence of aeration (\u003Ca href=\"#F2\"\u003EFigure 2\u003C\u002Fa\u003E). These stations were introduced in previous studies (\u003Ca href=\"#B28\"\u003EHarris et al., 2015\u003C\u002Fa\u003E). Water depths are between 1.5–3.5 m. In 2021, four upstream stations were added to determine the river influence to the aeration zone (\u003Ca href=\"#T2\"\u003ETable 2\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003ETABLE 1\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-t001.jpg\" name=\"Table1\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-t001.gif\" id=\"T1\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003ETABLE 1\u003C\u002Fstrong\u003E. Station information.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003ETABLE 2\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-t002.jpg\" name=\"Table2\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-t002.gif\" id=\"T2\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003ETABLE 2\u003C\u002Fstrong\u003E. Overview of manipulation experiments.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"mb0\"\u003ETo study the effects of the aeration, our experimental approach was to sample while the aerators had been on for about 1 month, referred to as the “ON” sampling event, which occurred during the daylight hours. In the evening of the “ON” sampling, the aerators were turned off and waters sampled 1 day later (2016; large bubble aeration), 7 days later (2018; large bubble aeration), 13 days later (2019; small bubble aeration), and 3 days later (2021; small bubble aeration); these are referred to as the “OFF” sampling events (\u003Ca href=\"#T2\"\u003ETable 2\u003C\u002Fa\u003E). After completion of the experiments, the aerators were turned back ON for the remainder of the season. By sampling at different time periods after aerators were turned OFF, the experiment addressed the question of the impact aeration had on methane flux from the Rock Creek estuary.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E2.2 Sampling Description\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EDuring each field campaign, we collected water column hydrographic and chemical profiles, discrete water samples from the surface and bottom depths, and shallow (∼30 cm) sediment cores \u003Cem\u003Evia\u003C\u002Fem\u003E small boat. Water column temperature, salinity, and dissolved oxygen levels were recorded with a YSI EXO2 multiparameter sonde. In 2018, a benthic lander was placed at a dock (location shown as black cross in \u003Ca href=\"#F2\"\u003EFigure 2\u003C\u002Fa\u003E) at the edge of the aeration zone that included continuous temperature, salinity, and dissolved oxygen sensors (YSI EXO2). Salinity is reported in practical salinity scale which has no units. Wind speeds were determined from a handheld anemometer (Weatherhawk, Windmate WM-200).\u003C\u002Fp\u003E\u003Ch4\u003E2.2.1. Discrete Water Column Samples\u003C\u002Fh4\u003E\u003Cp class=\"mb0\"\u003EWater samples were collected for dissolved methane when the aerators were ON and OFF. Water samples were always taken within 1 m of the same GPS location. The sampling location is ∼3 m off-axis to the aerators to ensure to not entangle the boat anchor. Therefore, when the aerators were ON, sampling never occurred in the bubble plume itself, always several meters away. Water column samples for dissolved methane concentrations were collected using published methods (\u003Ca href=\"#B48\"\u003EMagen et al., 2014\u003C\u002Fa\u003E). Briefly, a submersible pump was placed at either 50 cm from the air-water interface, or 1 m from the sediment bottom and dispensed water into 125 mL glass serum vials by overfilling 5 times the vial volume and avoiding bubbles. The vials were then capped with thick butyl rubber septa and crimp sealed with aluminum rings. A 10 mL air (Ultra Zero Air purity, Airgas) headspace was given to the vials and then 0.5 mL 8 M KOH was added to arrest microbial activity during storage. The samples were stored upside down at 4°C until they could be measured back at the laboratory with a headspace equilibration technique. In 2021, water was collected with a slight modification to the method where the headspace equilibration step was conducted \u003Cem\u003Ein situ\u003C\u002Fem\u003E and then the headspace physically separated from the water sample so no preservation was needed. Briefly, the submersible pump filled 120 mL into a 140 mL plastic syringe, bubble free. Then, 20 mL air (Ultra Zero Air purity, Airgas) was added and shook for 4 min to equilibrate. The temperature of the water was recorded for solubility calculations. Since this was a modification, we conducted efficiency tests using lab standards prior to the field campaign to verify 100% methane recovery from the method (\u003Ca href=\"#SM1\"\u003ESupplementary Figure S1\u003C\u002Fa\u003E). A complimentary water sample was also collected to quantify dissolved inorganic nitrogen (DIN) concentrations using published methods (\u003Ca href=\"#B28\"\u003EHarris et al., 2015\u003C\u002Fa\u003E), in all years but 2021.\u003C\u002Fp\u003E\u003Ch4\u003E2.2.2. Sediments\u003C\u002Fh4\u003E\u003Cp class=\"mb0\"\u003ESediment cores were collected by hand off the side of the boat using a 6.5 cm diameter plexiglass cylinder attached to a pole. The cores were brought back to shore and immediately (within 1 h of sampling) sliced into 3 cm vertical sections and the sediment packed into 50 mL centrifuge tubes and stored at 4°C for later analysis of pore-water sulfate concentrations (\u003Ca href=\"#B42\"\u003ELapham et al., 2008b\u003C\u002Fa\u003E). A separate sample for dissolved methane in the pore-waters was also collected by subcoring each section with a 3 mL cut off plastic syringe and placing the material in a 13.5 mL glass serum vial, capped with butyl rubber septae and preserved with 3 mL 1 M KOH (\u003Ca href=\"#B42\"\u003ELapham et al., 2008b\u003C\u002Fa\u003E). Sediment samples were stored at −20°C until analysis.\u003C\u002Fp\u003E\u003Ch4\u003E2.2.3. Air Samples\u003C\u002Fh4\u003E\u003Cp class=\"mb0\"\u003EAt each station, a 140 mL plastic syringe was used to collect an air sample above the sampling site. In 2018, more air samples were collected over time because of opportunistic sampling. Notably, on the ON and OFF days, air samples were collected at all stations at dawn, during the day, and at dusk. Sampling in 2019 was synchronous with water and sediment sampling during the day only. The syringe was upwind of any boat traffic and flushed copiously to provide a clean sample, before any other sampling occurred and potentially contaminated the air. These air samples were stored at 23°C for less than 2 days before they were measured for methane concentrations and stable carbon isotope ratios.\u003C\u002Fp\u003E\u003Ch4\u003E2.2.4. Continuous Bottom Water Sampling\u003C\u002Fh4\u003E\u003Cp class=\"mb0\"\u003ETo capture the temporal variability of methane concentrations between sampling campaigns and after the aeration manipulation experiments were completed, bottom water was continuously collected using OsmoSamplers at RC1, RC2, and RC7. OsmoSamplers (\u003Ca href=\"#SM1\"\u003ESupplementary Figures S2A,B\u003C\u002Fa\u003E) are osmotically-driven pumps that continuously collect and store water in narrow bore copper capillary tubing (\u003Ca href=\"#B33\"\u003EJannasch et al., 2004\u003C\u002Fa\u003E). They have been used in numerous natural environments to quantify dissolved methane concentrations, including from deep water methane seeps (\u003Ca href=\"#B41\"\u003ELapham et al., 2008a\u003C\u002Fa\u003E; \u003Ca href=\"#B83\"\u003EWilson et al., 2015\u003C\u002Fa\u003E), estuaries (\u003Ca href=\"#B24\"\u003EGelesh et al., 2016\u003C\u002Fa\u003E), high altitude rivers (\u003Ca href=\"#B9\"\u003EBuser-Young et al., 2021\u003C\u002Fa\u003E), high latitude wetlands (\u003Ca href=\"#B10\"\u003EBuser-Young et al., 2022\u003C\u002Fa\u003E) and arctic lakes (\u003Ca href=\"#B57\"\u003EMcIntosh Marcek et al., 2021\u003C\u002Fa\u003E). Osmosis in the pumps is created by an osmotic potential between a saturated brine chamber and freshwater chamber separated by semi-permeable membranes; no power is needed and there are no moving parts (\u003Ca href=\"#B79\"\u003ETheeuwes and Yum, 1976\u003C\u002Fa\u003E; \u003Ca href=\"#B33\"\u003EJannasch et al., 2004\u003C\u002Fa\u003E). The osmotic pump is then connected to small-bore (0.082 cm inner diameter), long (up to 300 m) copper tubing coil, that is prefilled with freshwater. Water is then continuously drawn from the end of the copper tubing and stored in this tubing over time. The copper material is used so gases (i.e., methane) do not diffuse through it. The pumping rate is positively correlated to the surrounding water temperature and the number of membranes in the pump. For these deployments, two pump speeds were used. “Slow” pumps (8 membranes which pump 1 mL day\u003Csup\u003E−1\u003C\u002Fsup\u003E) were used for the long term collection of bottom water through the summer to give a temporal resolution of about 5 days (deployed for 9 months). “Fast” pumps (44 membranes which pump 5 mL day\u003Csup\u003E−1\u003C\u002Fsup\u003E) were used for a temporal high resolution of ∼1 day (deployed for 1 week). For the slow pumps, we assume the sample stream undergoes plug flow; thus, dispersion within the tubing is minimal (\u003Ca href=\"#B33\"\u003EJannasch et al., 2004\u003C\u002Fa\u003E). For the fast pumps, plug flow may not be met, so we limited the deployment to a week. OsmoSamplers only collect water thereby precluding gas from affecting the sampler. The intakes are fitted with a 0.2 µm rhizone filter (\u003Ca href=\"#B74\"\u003ESeeberg-Elverfeldt et al., 2005\u003C\u002Fa\u003E) to preclude microbes from the collection and alter the sample stream in the tubing during the deployment.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EAt the time of the deployment, the OsmoSamplers were attached to the copper coils and placed in a plastic crate (33 × 33 × 28 cm) under 16 kg weight and tied to a surface buoy (\u003Ca href=\"#SM1\"\u003ESupplementary Figures S2C–E\u003C\u002Fa\u003E). In 2018, the fast OsmoSampler sets were deployed at stations RC1, RC2, and RC7. In 2018, and 2019, slow OsmoSampler sets were also deployed at each of those stations along with Onset temperature and conductivity loggers. Because Rock Creek is a dynamic estuary, we used the conductivity detectors to verify the time-stamps in the OsmoSampler coils. OsmoSamplers were deployed from 9 July to 9 October 2018 and 18 June 2019 to 28 October 2019.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EUpon recovery, the copper coils were sealed on either end with pliers and taken back to the lab to be stored at 4°C prior to further processing in the lab. The sensor data were downloaded. Within 1 week, the copper coils were unspooled and crimped into alternating lengths of 50 cm and 4.5 m using a wire crimping tool (\u003Ca href=\"#B24\"\u003EGelesh et al., 2016\u003C\u002Fa\u003E). The 50 cm sections were squeezed with a bench-top roller to flatten the copper tubing and force the liquid into 2 mL plastic tubes to immediately test for salinity using a handheld Extech RF20 refractometer. Because the coils were prefilled with freshwater before deployment, sectioning was terminated when zero salinity was observed for three samples in a row. These samples were also measured for chloride concentrations to compare to sensor conductivity measurements to verify time-stamps (\u003Ca href=\"#B24\"\u003EGelesh et al., 2016\u003C\u002Fa\u003E). The 4.5 m copper coil sections were squeezed using the bench-top hand roller which expressed sample liquid through a gastight adaptor and needle, and into a 13.5 mL glass sample vial at the opposite end, previously capped with a butyl rubber septum to prevent gas exchange, and flushed with helium. Each 4.5 m copper section contained approximately 2 mL of sample liquid that was transferred to the vials, resulting in an initial overpressure of approximately 2 mL. The dissolved CH\u003Csub\u003E4\u003C\u002Fsub\u003E equilibrated with the helium headspace after shaking the vial for 2 min. Time stamps were calculated by adjusting pumping rates to \u003Cem\u003Ein situ\u003C\u002Fem\u003E temperature, as shown in \u003Ca href=\"#B24\"\u003EGelesh et al. (2016)\u003C\u002Fa\u003E. Unfortunately, the fast pumps only had 1 weeks’ worth of tubing yet were deployed for 12 days due to weather delays. Thus, the pumps overpumped the coil and the first part of the deployment was lost.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E2.3 Analytical Methods\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EFor water column and sediment vial samples, the headspace was equilibrated with the dissolved methane from the aqueous sample and the headspace analyzed for the ppmv methane. For all samples, an aliquot of the headspace was extracted from the vials and injected onto a gas chromatograph (SRI 8610C multi-gas) equipped with a HayeSep D packed column and a Flame Ionization Detector in order to quantify methane concentrations (\u003Ca href=\"#B48\"\u003EMagen et al., 2014\u003C\u002Fa\u003E). Certified standards (Airgas, Inc.) were used for the calibration curve. Analytical precision is 3% and all measurements were above detection limit of 2 ppmv. Resultant partial pressures were then used to calculate dissolved methane concentrations (in nM) in either the water column or the porewater using Henry’s law according to equations in \u003Ca href=\"#B48\"\u003EMagen et al. (2014)\u003C\u002Fa\u003E and porosity corrections according to \u003Ca href=\"#B41\"\u003ELapham et al. (2008a)\u003C\u002Fa\u003E.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EFor sulfate and chloride concentrations in the sediment porewater samples, the tube containing whole sediment was centrifuged (3000 RPM, 30 min, 20°C, Sorvall\u003Csup\u003E©\u003C\u002Fsup\u003E RT 6000D) and the resultant supernatant filtered with a 0.2 µm syringe filter. Samples were then diluted (1:135) in Milli-Q water and analyzed on a Dionex ICS 1000 ion chromatograph (IonPac AG22 4 × 50 mm guard column, IonPac AS22 4 × 250 mm analytical column, and ASRS 300 4 mm suppressor) with an AS40 Autosampler. Water samples from the 50 cm OsmoSampler sections were also measured for chloride with the same dilution to calculate salinity. Certified IAPSO seawater standard (Ocean Scientific International Ltd.) was used for the calibration curve. Analytical precision is 2% and all measurements were above detection limit of 0.05 mM for sulfate.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EAir syringes were directly connected to the intake of a cavity ring down spectrometer (CRDS, Picarro 2201i) to measure for methane concentrations and methane stable carbon isotopes. For the water samples, 10 mL of degassed brine was added to the vials to displace the headspace and injected into the small sample isotope module (Picarro, Inc.) to introduce a small sample to the CRDS, similar to procedure in \u003Ca href=\"#B57\"\u003EMcIntosh Marcek et al. (2021)\u003C\u002Fa\u003E. Isotope values were obtained through calibration with three Vienna Pee Dee Belemnite (VPDB) referenced standards (−23.9‰, −38.3‰, and −66.5‰ (±0.2‰); Isometric Instruments). Isotopic results are reported using the δ\u003Csup\u003E13\u003C\u002Fsup\u003EC notation in per mil (‰), where δ\u003Csup\u003E13\u003C\u002Fsup\u003EC = (R\u003Csub\u003Esample\u003C\u002Fsub\u003E\u002FR\u003Csub\u003Estandard\u003C\u002Fsub\u003E -1)*1,000 and R = \u003Csup\u003E13\u003C\u002Fsup\u003EC\u002F\u003Csup\u003E12\u003C\u002Fsup\u003EC. Analytical precision is 2% for concentrations and 4‰ for stable carbon isotope ratios.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E2.4 Calculations\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EThe air-water flux of CH\u003Csub\u003E4\u003C\u002Fsub\u003E, F, was determined for all discrete sampling campaigns using the updated flux equations presented \u003Ca href=\"#B80\"\u003EWanninkhof (2014)\u003C\u002Fa\u003E:\u003C\u002Fp\u003E\u003Cdiv class=\"equationImageholder\"\u003E\u003Cmath id=\"e1\"\u003E\u003Cmrow\u003E\u003Cmi\u003EF\u003C\u002Fmi\u003E\u003Cmo\u003E=\u003C\u002Fmo\u003E\u003Cmi\u003Ek\u003C\u002Fmi\u003E\u003Cmrow\u003E\u003Cmo\u003E(\u003C\u002Fmo\u003E\u003Cmrow\u003E\u003Cmsub\u003E\u003Cmi\u003EC\u003C\u002Fmi\u003E\u003Cmi\u003Ew\u003C\u002Fmi\u003E\u003C\u002Fmsub\u003E\u003Cmo\u003E−\u003C\u002Fmo\u003E\u003Cmsub\u003E\u003Cmi\u003EC\u003C\u002Fmi\u003E\u003Cmrow\u003E\u003Cmi\u003Ee\u003C\u002Fmi\u003E\u003Cmi\u003Eq\u003C\u002Fmi\u003E\u003C\u002Fmrow\u003E\u003C\u002Fmsub\u003E\u003C\u002Fmrow\u003E\u003Cmo\u003E)\u003C\u002Fmo\u003E\u003C\u002Fmrow\u003E\u003C\u002Fmrow\u003E\u003Cmspace width=\"5em\"\u002F\u003E\u003Cmo stretchy='false'\u003E(\u003C\u002Fmo\u003E\u003Cmn\u003E1\u003C\u002Fmn\u003E\u003Cmo stretchy='false'\u003E)\u003C\u002Fmo\u003E\u003C\u002Fmath\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"noindent\"\u003Ewhere k is the gas transfer velocity (length time\u003Csup\u003E−1\u003C\u002Fsup\u003E), C\u003Csub\u003Ew\u003C\u002Fsub\u003E is the measured surface water concentration, and C\u003Csub\u003Eeq\u003C\u002Fsub\u003E is the CH\u003Csub\u003E4\u003C\u002Fsub\u003E concentration in equilibrium with the atmosphere at \u003Cem\u003Ein situ\u003C\u002Fem\u003E conditions (\u003Ca href=\"#B84\"\u003EYamamoto et al., 1976\u003C\u002Fa\u003E). There are several versions of \u003Ca href=\"#e1\"\u003EEq. 1\u003C\u002Fa\u003E, mostly based on wind speed. Here we employ the formulation and parameterization of \u003Ca href=\"#B61\"\u003EMyllykangas et al. (2020)\u003C\u002Fa\u003E, which reports a k value adapted from \u003Ca href=\"#B66\"\u003ERaymond and Cole (2001)\u003C\u002Fa\u003E:\u003C\u002Fp\u003E\u003Cdiv class=\"equationImageholder\"\u003E\u003Cmath id=\"e2\"\u003E\u003Cmrow\u003E\u003Cmi\u003Ek\u003C\u002Fmi\u003E\u003Cmo\u003E=\u003C\u002Fmo\u003E\u003Cmn\u003E1.91\u003C\u002Fmn\u003E\u003Cmo\u003E \u003C\u002Fmo\u003E\u003Cmsup\u003E\u003Cmi\u003Ee\u003C\u002Fmi\u003E\u003Cmrow\u003E\u003Cmn\u003E0.35\u003C\u002Fmn\u003E\u003Cmi\u003Eu\u003C\u002Fmi\u003E\u003C\u002Fmrow\u003E\u003C\u002Fmsup\u003E\u003Cmsup\u003E\u003Cmrow\u003E\u003Cmrow\u003E\u003Cmo\u003E(\u003C\u002Fmo\u003E\u003Cmrow\u003E\u003Cmfrac\u003E\u003Cmrow\u003E\u003Cmi\u003ES\u003C\u002Fmi\u003E\u003Cmi\u003Ec\u003C\u002Fmi\u003E\u003C\u002Fmrow\u003E\u003Cmrow\u003E\u003Cmn\u003E600\u003C\u002Fmn\u003E\u003C\u002Fmrow\u003E\u003C\u002Fmfrac\u003E\u003C\u002Fmrow\u003E\u003Cmo\u003E)\u003C\u002Fmo\u003E\u003C\u002Fmrow\u003E\u003C\u002Fmrow\u003E\u003Cmrow\u003E\u003Cmo\u003E−\u003C\u002Fmo\u003E\u003Cmn\u003E0.5\u003C\u002Fmn\u003E\u003C\u002Fmrow\u003E\u003C\u002Fmsup\u003E\u003C\u002Fmrow\u003E\u003Cmspace width=\"5em\"\u002F\u003E\u003Cmo stretchy='false'\u003E(\u003C\u002Fmo\u003E\u003Cmn\u003E2\u003C\u002Fmn\u003E\u003Cmo stretchy='false'\u003E)\u003C\u002Fmo\u003E\u003C\u002Fmath\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"noindent\"\u003Ewhere \u003Cem\u003Eu\u003C\u002Fem\u003E is the average wind speed and Sc is the Schmidt number for CH\u003Csub\u003E4\u003C\u002Fsub\u003E in freshwater calculated from \u003Ca href=\"#B80\"\u003EWanninkhof (2014)\u003C\u002Fa\u003E. Wind speeds were obtained from a nearby NOAA buoy (National Data Buoy Center BLTM2, Baltimore, MD) and averaged over the 3 days prior to sampling. Since the buoy is located 15 km away from Rock Creek, we compared the buoy to the handheld anemometer readings and found they compared within 7%. The buoy wind speed was used for all stations. Wind speeds varied between 2–3.2 m s\u003Csup\u003E−1\u003C\u002Fsup\u003E, which translated to k values varying between 3.8 and 5.9 cm h\u003Csup\u003E−1\u003C\u002Fsup\u003E, similar to values found in \u003Ca href=\"#B47\"\u003EMacIntyre et al. (2010)\u003C\u002Fa\u003E in a lake system and mangrove dominated estuaries (\u003Ca href=\"#B69\"\u003ERosentreter et al., 2017\u003C\u002Fa\u003E). Air-water fluxes were then calculated using \u003Ca href=\"#e1\"\u003EEq. 1\u003C\u002Fa\u003E and reported as µmol CH\u003Csub\u003E4\u003C\u002Fsub\u003E m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003ESince the calculated air-water flux is inherently based on assumptions of a stagnant boundary layer, the calculated air-water methane flux when the aerators are ON will be underestimated. To constrain this better when the aerators were ON, we calculated an air-water methane flux from direct bubble transport to surface water by applying the bubble radius of the system (3 mm, Mobley Engineering, Inc., personal communication) to an existing bubble model output to determine the mass transfer coefficient (\u003Ca href=\"#F4\"\u003EFigure 4\u003C\u002Fa\u003E in \u003Ca href=\"#B56\"\u003EMcGinnis and Little, 2002\u003C\u002Fa\u003E). This mass transfer coefficient for the bubble radius in the system is 0.04 cm s\u003Csup\u003E−1\u003C\u002Fsup\u003E (or 144 cm h\u003Csup\u003E−1\u003C\u002Fsup\u003E) which was then used in \u003Ca href=\"#e1\"\u003EEq. 1\u003C\u002Fa\u003E to calculate a modified air-water methane flux for 2018 and 2019 at RC1 and RC2.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EThe sediment-water methane diffusive flux was calculated from Fick’s first law:\u003C\u002Fp\u003E\u003Cdiv class=\"equationImageholder\"\u003E\u003Cmath id=\"e3\"\u003E\u003Cmrow\u003E\u003Cmsub\u003E\u003Cmi\u003EJ\u003C\u002Fmi\u003E\u003Cmrow\u003E\u003Cmi\u003EC\u003C\u002Fmi\u003E\u003Cmi\u003EH\u003C\u002Fmi\u003E\u003Cmn\u003E4\u003C\u002Fmn\u003E\u003Cmo\u003E−\u003C\u002Fmo\u003E\u003Cmi\u003ES\u003C\u002Fmi\u003E\u003Cmi\u003EW\u003C\u002Fmi\u003E\u003Cmi\u003EI\u003C\u002Fmi\u003E\u003C\u002Fmrow\u003E\u003C\u002Fmsub\u003E\u003Cmo\u003E=\u003C\u002Fmo\u003E\u003Cmo\u003E \u003C\u002Fmo\u003E\u003Cmo\u003E−\u003C\u002Fmo\u003E\u003Cmi\u003Eφ\u003C\u002Fmi\u003E\u003Cmsub\u003E\u003Cmi\u003ED\u003C\u002Fmi\u003E\u003Cmi\u003Es\u003C\u002Fmi\u003E\u003C\u002Fmsub\u003E\u003Cmfrac\u003E\u003Cmrow\u003E\u003Cmi\u003Ed\u003C\u002Fmi\u003E\u003Cmi\u003EC\u003C\u002Fmi\u003E\u003C\u002Fmrow\u003E\u003Cmrow\u003E\u003Cmi\u003Ed\u003C\u002Fmi\u003E\u003Cmi\u003Ex\u003C\u002Fmi\u003E\u003C\u002Fmrow\u003E\u003C\u002Fmfrac\u003E\u003C\u002Fmrow\u003E\u003Cmspace width=\"5em\"\u002F\u003E\u003Cmo stretchy='false'\u003E(\u003C\u002Fmo\u003E\u003Cmn\u003E3\u003C\u002Fmn\u003E\u003Cmo stretchy='false'\u003E)\u003C\u002Fmo\u003E\u003C\u002Fmath\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"noindent\"\u003Ewhere J\u003Csub\u003ECH4-SWI\u003C\u002Fsub\u003E is the methane flux (µmol CH\u003Csub\u003E4\u003C\u002Fsub\u003E cm\u003Csup\u003E−2\u003C\u002Fsup\u003E yr\u003Csup\u003E−1\u003C\u002Fsup\u003E) at the sediment-water interface, φ is the porosity (0.8), D\u003Csub\u003Es\u003C\u002Fsub\u003E is the sedimentary methane diffusion coefficient (cm\u003Csup\u003E2\u003C\u002Fsup\u003E s\u003Csup\u003E−1\u003C\u002Fsup\u003E), x is the vertical sediment depth (cm) and dC\u002Fdx is the concentration gradient of methane. D\u003Csub\u003Es\u003C\u002Fsub\u003E was calculated for each station, corrected for tortuosity and \u003Cem\u003Ein situ\u003C\u002Fem\u003E pressures (based on water depth), temperatures, and salinity (\u003Ca href=\"#B59\"\u003EMillero, 1996\u003C\u002Fa\u003E), and was ∼1.7 × 10\u003Csup\u003E–6\u003C\u002Fsup\u003E cm\u003Csup\u003E2\u003C\u002Fsup\u003E s\u003Csup\u003E−1\u003C\u002Fsup\u003E. The gradient term was calculated between the uppermost porewater measurements; which usually started at 1.5 cm into the sediment. Thus, this gradient is most likely overestimated because it ignores any oxidation processes that might occur in that upper 1.5 cm of sediment. Using this gradient, the diffusive fluxes were then calculated for each station and time point with \u003Ca href=\"#e3\"\u003EEq. 3\u003C\u002Fa\u003E. For convenience, fluxes are reported as positive but represent flux out of sediment).\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E2.5 Box Model\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EWe applied a simple box model to the flux data with the assumption that the sediments are the only source of methane to the water, and the atmospheric flux was the only sink of methane in the water. If the two balanced, then there would be no additional contributions to the methane budget in Rock Creek. If the fluxes did not balance, we could invoke additional microbial oxidation or production in the water column and\u002For advective transport of methane from up- or down-stream. To do this, at each station for 2018 and 2019, we subtracted the atmospheric methane flux from the sedimentary methane flux, then assigned the net difference as the water column methane inventory anomaly. For the aeration sites (RC1 and RC2), both air-water and sedimentary fluxes were averaged together. We should note that when the aerators are ON, we used the fluxes estimated from the stagnant boundary layer model which most likely underestimates the atmospheric flux.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E2.6 Statistics\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EStudent t-test was used in Excel (two-tailed, paired) to determine if the methane concentrations calculated from the discrete sampling events and the OsmoSampler samples were significantly similar. To create the paired methane dataset for this test, we extracted the OsmoSampler methane concentration that was sampled at the same time as the discrete methane measurement from the water column at each station. Please note that the OsmoSamplers average over 6–12 h. This comparison was done with the fast OsmoSamplers during 2018 (see Section 3.6 for result).\u003C\u002Fp\u003E\u003Ca id=\"h4\" name=\"h4\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003E3 Results\u003C\u002Fh2\u003E\u003Ch3 class=\"pt0\"\u003E3.1 Water Column Oxygen, Dissolved Inorganic Nitrogen, Temperature, and Salinity\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EThe experimental design was intended to measure methane concentrations in well oxygenated, bubble-influenced waters when aerators were ON, and then hypoxic or anoxic waters without bubble transport when the aerators were turned OFF. Based on previous experiments in Rock Creek, the change to hypoxic conditions occurred within 1 day of turning off the aerators (\u003Ca href=\"#B28\"\u003EHarris et al., 2015\u003C\u002Fa\u003E). However, for our manipulations, the goal was to observe changes over the longer term (up to 13 days) and the systems response. It is important to remember there was a 7-day difference between ON and OFF treatments in 2018 and 13-day difference in 2019.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EIn 2018 during aeration, the dissolved oxygen (DO) concentrations were ∼7 mg L\u003Csup\u003E−1\u003C\u002Fsup\u003E at RC1 and RC2, and well mixed (\u003Ca href=\"#F3\"\u003EFigure 3\u003C\u002Fa\u003E top panel). At RC7, the water column was still oxygenated (>4 mg L\u003Csup\u003E−1\u003C\u002Fsup\u003E; comparable to conditions at RC9b). After aerators were OFF for 7 days, RC1 and RC2 became hypoxic throughout the water column below 0.5 m, and the other stations had weakly stratified water columns, but with hypoxic bottom waters below 2.5 m. Dissolved inorganic nitrogen concentrations averaged 14.4 µM (ranging between 0.8 and 19 µM for all stations) when aerators were ON and 7.2 µM (ranging between 3.7 and 12 µM) when aerators were OFF (\u003Ca href=\"#T2\"\u003ETable 2\u003C\u002Fa\u003E). Salinity was ∼5–5.5 (\u003Ca href=\"#F3\"\u003EFigure 3\u003C\u002Fa\u003E). Water temperatures in 2018 varied between 26–30°C, and were similar for both ON and OFF treatments (data not shown).\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 3\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g003.jpg\" name=\"Figure3\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g003.gif\" id=\"F3\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 3\u003C\u002Fstrong\u003E. Water column salinity and dissolved oxygen (O\u003Csub\u003E2\u003C\u002Fsub\u003E) profiles from 2018 (top panels) and 2019 (bottom panels) for stations within the aeration zone and stations outside the aeration zone. Blue symbols signify when aerators were ON and red when they were OFF.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"mb0\"\u003EIn 2019, dissolved oxygen was relatively high in surface waters when the aerators were OFF at RC1 and RC2, giving way to oxygen-depleted conditions below 2 m (\u003Ca href=\"#F3\"\u003EFigure 3\u003C\u002Fa\u003E). Dissolved inorganic nitrogen concentrations averaged 27.8 µM (ranging between 5 and 36 µM for all stations) when aerators were ON and 4.9 µM (ranging between 1 and 13 µM) when aerators were OFF (\u003Ca href=\"#T2\"\u003ETable 2\u003C\u002Fa\u003E). Salinity was lower in 2019 than in 2018 and was nearly 3.5 when aerators were ON, and 4.5 when aerators were OFF (\u003Ca href=\"#F3\"\u003EFigure 3\u003C\u002Fa\u003E). Water temperatures were between 26 and 29°C during the ON treatment and were warmer during OFF treatment (data not shown).\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EThe tidal stage at each station varied over the discrete sampling time points (\u003Ca href=\"#SM1\"\u003ESupplementary Figure S3\u003C\u002Fa\u003E). In 2018 when aerators were ON, RC7 was sampled first at the ebbing tide, RC9b at low tide, and RC1 and RC2 at a high tide. When aerators were OFF, RC1 and RC2 were collected close to high tide or when waters were just beginning to ebb. RC7 and RC9b were sampled on the ebb tide. In 2019 when aerators were ON, RC9b, RC1, and RC2 were collected on flooding tide, and RC7 was collected right after high tide. When aerators were OFF, all stations were collected near the high tide.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E3.2 Methane Concentrations and Stable Carbon Isotope Ratios in Water\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EDissolved methane concentrations in surface and bottom water of Rock Creek varied over space and time (\u003Ca href=\"#F4\"\u003EFigure 4\u003C\u002Fa\u003E). Overall, concentrations ranged between 150 and 1,500 nM, orders of magnitude higher than atmospheric equilibrium (which is ∼3 nM), and were higher at stations RC1 and RC2, within the aeration zone, compared to stations closer to the Patapsco River (RC7, RC9b). In 2016, concentrations at RC2 and RC7 were around 400 nM throughout the period of measurements, regardless of aeration status (\u003Ca href=\"#F4\"\u003EFigure 4A\u003C\u002Fa\u003E). In 2018, during aeration, the waters were relatively well mixed between surface and bottom waters (\u003Ca href=\"#F4\"\u003EFigure 4B\u003C\u002Fa\u003E). Once aerators were turned OFF, there was an increase in bottom water methane at RC1 and RC2, and not much change at RC7 and RC9b. After 7 days, the surface waters were enriched in methane compared to the bottom water. In 2019, the concentrations between surface and bottom water followed expectations: during aeration, the water column was well mixed so there was little difference between surface and bottom waters and when aerators were turned OFF, methane concentrations were higher in the bottom water than surface waters after 13 days (\u003Ca href=\"#F4\"\u003EFigure 4C\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 4\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g004.jpg\" name=\"Figure4\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g004.gif\" id=\"F4\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 4\u003C\u002Fstrong\u003E. Methane concentrations (colored bars) in surface (S) and bottom (B) water at all stations in \u003Cstrong\u003E(A)\u003C\u002Fstrong\u003E 2016, \u003Cstrong\u003E(B)\u003C\u002Fstrong\u003E 2018, \u003Cstrong\u003E(C)\u003C\u002Fstrong\u003E 2019, and \u003Cstrong\u003E(D)\u003C\u002Fstrong\u003E 2021. All blue colors represent the “ON” situation, and the gradients in gray color represent the number of days after aerators were turned off, which are described in each panel. Error bars represent standard error on replicate samples collected.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"mb0\"\u003EThe 2021 field campaign was designed to resolve the upstream contribution of methane to the aeration zone. Overall, methane concentrations were lower than 2019 values but also showed the same pattern of higher concentrations at RC1 and RC2, than at RC7 (\u003Ca href=\"#F4\"\u003EFigure 4D\u003C\u002Fa\u003E). At RC7, concentrations were similar between surface and bottom, and regardless of aeration status. At RC1 and RC2, methane concentrations were lower when aerators were OFF, which was unexpected. The other unexpected result was to record higher methane concentrations in surface waters than bottom waters during both ON and OFF periods (\u003Ca href=\"#F4\"\u003EFigure 4D\u003C\u002Fa\u003E). Measurements made upstream of the aerators, which were only made in 2021, showed that surface waters were always higher than the bottom waters and that concentrations during the ON treatment were always higher than the OFF. Furthermore, methane concentrations were highest in the most upstream station, and declined downstream and into the estuary, such that water flowing into the aeration zone from upstream were enriched with methane relative to the aeration zone itself.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EThe water column methane concentrations are also presented in \u003Ca href=\"#F5\"\u003EFigures 5A–D\u003C\u002Fa\u003E, \u003Ca href=\"#F6\"\u003E6A–D\u003C\u002Fa\u003E as compilation figures showing the water column and sediments in a holistic view. Here we add to the water column concentration data the stable isotopic ratio of methane carbon to distinguish source of this methane (\u003Ca href=\"#F5\"\u003EFigures 5I–L\u003C\u002Fa\u003E, \u003Ca href=\"#F6\"\u003E6I–L\u003C\u002Fa\u003E). Regardless of station or aeration status, δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values ranged from −69 to −51‰, with an average value of −61.4 ± 3.6‰, which is near the standard deviation of the method.\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 5\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g005.jpg\" name=\"Figure5\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g005.gif\" id=\"F5\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 5\u003C\u002Fstrong\u003E. In 2018, methane concentrations \u003Cstrong\u003E(A–H)\u003C\u002Fstrong\u003E and stable carbon isotopes \u003Cstrong\u003E(I–O)\u003C\u002Fstrong\u003E in water column (blue background) and in sediments (brown background). Blue symbol color denotes when aerators were on, and dark gray when they were off for 7 days. Horizontal bars in sediments \u003Cstrong\u003E(M–O)\u003C\u002Fstrong\u003E shows the movement of the sulfate-methane transition zone between ON (blue color) and OFF (gray color).\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 6\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g006.jpg\" name=\"Figure6\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g006.gif\" id=\"F6\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 6\u003C\u002Fstrong\u003E. In 2019, methane concentrations \u003Cstrong\u003E(A–H)\u003C\u002Fstrong\u003E and stable carbon isotopes \u003Cstrong\u003E(I–O)\u003C\u002Fstrong\u003E in water column (blue background) and in sediments (brown background). Blue symbol color denotes when aerators were on, and dark gray when they were off for 13 days. Horizontal bars in sediments \u003Cstrong\u003E(M–O)\u003C\u002Fstrong\u003E shows the movement of the sulfate-methane transition zone between ON (blue color) and OFF (gray color).\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Ch3 class=\"pt0\"\u003E3.3 Sediment Porewater\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EMethane concentrations measured from the sediment porewaters increased with sediment depth (\u003Ca href=\"#F5\"\u003EFigures 5\u003C\u002Fa\u003E, \u003Ca href=\"#F6\"\u003E6\u003C\u002Fa\u003E, brown colored panels). Surface concentrations were at ∼µM levels and increased to as high as 1,300 µM at the bottom of the core. In both years, concentrations were higher in the aeration zone (RC1 and RC2) and outside the aeration zone (RC7) compared to the background site (RC9b). Yet, there is variability in the sediment profiles. For example, at RC1 and RC2, sediment methane concentrations were lower in 2018 than in 2019, regardless of aeration status. After 13 days of no aeration, methane concentrations were higher in the sediments at RC2 and RC7. As noted, \u003Ca href=\"#F5\"\u003EFigures 5\u003C\u002Fa\u003E, \u003Ca href=\"#F6\"\u003E6\u003C\u002Fa\u003E also contain methane water column parameters for comparison purposes.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EMethane increases in sediment porewaters are typically associated with a drawdown of sulfate in the surficial depths due to sulfate reducers outcompeting methanogens for substrates (\u003Ca href=\"#B31\"\u003EHoehler et al., 1994\u003C\u002Fa\u003E). Therefore, we also measured sulfate in porewater to help our understanding of anaerobic biogeochemical processing. Sulfate concentrations decreased downcore in all stations except RC9b although the depth of low sulfate (SO\u003Csub\u003E4\u003C\u002Fsub\u003E < 0.5 mM) varied (\u003Ca href=\"#SM1\"\u003ESupplementary Figure S4\u003C\u002Fa\u003E). The depth of low sulfate typically coincides with the increase of methane, and is known as the sulfate methane transition (SMT) depth. The SMT is an area of active anaerobic methane oxidation \u003Cem\u003Evia\u003C\u002Fem\u003E sulfate reduction (\u003Ca href=\"#B36\"\u003EJørgensen et al., 2020\u003C\u002Fa\u003E) and is a useful metric to show how active the anaerobic microbial community is in a sediment column (\u003Ca href=\"#F5\"\u003EFigures 5M–O\u003C\u002Fa\u003E, \u003Ca href=\"#F6\"\u003E6M–O\u003C\u002Fa\u003E). While we expected to see the SMT depth shoal when aeration was turned OFF, there was no consistent pattern of the depth of the SMT with aeration status for both years, although we will specifically present SMT depth patterns in each year below. Using the gradients from the top of the cores, the methane flux to the sediment-water interface varied across space and time, and ranged between 0.1 and 700 µmol m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E (\u003Ca href=\"#SM1\"\u003ESupplementary Figure S5\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EChloride concentrations were also measured as a conservative tracer and as a way to validate any depletion of sulfate coming from sulfate reduction and not a consequence of groundwater. In 2018, while the chloride concentrations showed a slight increase in depth, the depth averages are as follows: 76 ± 8 mM (RC1, RC2), 100 ± 14 mM (RC7), and 85 ± 10 mM (RC9). Since these chloride values are within the range of what would be expected given the overlying water salinity, we represent any conservative mixing in terms of how it might affect sulfate concentrations. Given the rule of constant proportions, we calculated the range of sulfate values that would be estimated given those chloride concentrations (shaded rectangles in \u003Ca href=\"#SM1\"\u003ESupplementary Figure S4\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EMethane stable carbon isotope ratios were measured to help distinguish the fate of methane formed in the sediment. Our measurements revealed two patterns in δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values with depth: 1) δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E increasing with depth and 2) δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E peaks at intermediate depths associated with the SMT. In 2018, at RC1, the δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values were similar during the ON and OFF conditions in that the surface was \u003Csup\u003E13\u003C\u002Fsup\u003EC depleted (between −80 and −70‰), they became heavier with depth to as high as −50‰, and then decreased again to near surface sediment values (\u003Ca href=\"#F5\"\u003EFigure 5M\u003C\u002Fa\u003E). The depth of the SMT deepened with aerators OFF. At RC2, during the ON condition, δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values were ∼−80‰ and increased with depth in the core (\u003Ca href=\"#F5\"\u003EFigure 5N\u003C\u002Fa\u003E). During the OFF condition, δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values were around −60‰ in the shallow depths and quickly decreased to −80‰ at the SMT. At RC7, just outside the aeration zone, δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values were ∼−70‰ at the surface when waters were aerated and decreased to −85‰ at the bottom of the core (\u003Ca href=\"#F5\"\u003EFigure 5O\u003C\u002Fa\u003E). During the OFF situation, values showed a similar trend at the surface but then showed a mid-depth minimum of −80‰ at the SMT.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EIn 2019, the δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E value trends showed more consistency across the stations (\u003Ca href=\"#F6\"\u003EFigures 6M–O\u003C\u002Fa\u003E). At RC1, values decreased at the surface from as high as −40 to ∼ −70‰ where values remained for about 15 cm into the sediments (\u003Ca href=\"#F6\"\u003EFigure 6M\u003C\u002Fa\u003E). RC2 and RC7 show almost the same isotope profiles where values decrease downcore, but the δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values are ∼10–15‰ higher when aerators were ON compared to when they were OFF (\u003Ca href=\"#F6\"\u003EFigures 6N,O\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E3.4 Methane Concentrations and Stable Carbon Isotope Ratios in Air and Air-Water Fluxes\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EMethane concentrations were measured in the air above each station during 2018 and 2019. In 2018, the average atmospheric methane concentration across all sites was 1.84 ± 0.06 ppmv, and didn’t vary between ON and OFF conditions, except at the dawn sampling (\u003Ca href=\"#F7\"\u003EFigures 7A,B\u003C\u002Fa\u003E). The average δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E value of the background methane was −51.4 ± 10‰ (\u003Ca href=\"#F7\"\u003EFigures 7C,D\u003C\u002Fa\u003E). Dawn sampling on the ON and OFF days showed elevated methane concentrations, reaching as high as 3 ppmv at RC7 which had a δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E value of −90‰ (\u003Ca href=\"#F7\"\u003EFigures 7A,B\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 7\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g007.jpg\" name=\"Figure7\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g007.gif\" id=\"F7\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 7\u003C\u002Fstrong\u003E. Methane concentrations in air above the water in \u003Cstrong\u003E(A)\u003C\u002Fstrong\u003E 2018 and \u003Cstrong\u003E(B)\u003C\u002Fstrong\u003E 2019. δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values in air above the water in \u003Cstrong\u003E(C)\u003C\u002Fstrong\u003E 2018 and \u003Cstrong\u003E(D)\u003C\u002Fstrong\u003E 2019. Shaded regions indicates when aerators were OFF.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"mb0\"\u003ETo calculate the air-water methane flux, we used two approaches. The first used the stagnant boundary layer model and most likely underestimates the flux for when the aerators are ON. Using this model, the air-water methane fluxes ranged between 300 and 1,500 µmol CH\u003Csub\u003E4\u003C\u002Fsub\u003E m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E (\u003Ca href=\"#F8\"\u003EFigure 8\u003C\u002Fa\u003E). The flux was higher at RC1 and RC2 than other stations, regardless of aeration status or year. In 2018, the flux at the RC1 and RC2 was higher when aerators were ON after 7 days, whereas in 2019, the flux was lower when the aerators were ON. The second approach was only carried out when the aerators were ON and assumed methane was being stripped from the water as the aerator bubbles traveled up the water column. The calculated fluxes were much higher than the fluxes from the stagnant boundary layer (\u003Ca href=\"#F8\"\u003EFigure 8\u003C\u002Fa\u003E extended arrows to blue dots). In 2018, at RC1 and RC2, air-water methane flux was 30,730 and 19,380 µmol CH\u003Csub\u003E4\u003C\u002Fsub\u003E m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E, respectively. In 2019, at RC1 and RC2, air-water methane flux was 14,669 and 8,342 µmol CH\u003Csub\u003E4\u003C\u002Fsub\u003E m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E, respectively.\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 8\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g008.jpg\" name=\"Figure8\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g008.gif\" id=\"F8\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 8\u003C\u002Fstrong\u003E. Air-water methane flux for 2018 and 2019. Aerated waters are in blue, and non-aerated waters are shown in gray scale that corresponds to the number of days aerators were off. The bars indicate the flux calculated with the stagnant boundary layer model, whereas the extended arrows to the blue dots indicate the flux recalculated with bubble influence.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Ch3 class=\"pt0\"\u003E3.5 Box Model Results\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EThe sediment and air-water fluxes were used in the box model to determine if there are additional sources or sinks of methane beyond what is coming from the sediments and being lost to the atmosphere (\u003Ca href=\"#F9\"\u003EFigure 9\u003C\u002Fa\u003E). In 2018, during large bubble aeration, there was a large source of methane (positive values in \u003Ca href=\"#F9\"\u003EFigure 9\u003C\u002Fa\u003E) at the aerators, and actually a methane sink from RC7 when the aerators were off. In 2019, there was a methane source across all stations, regardless of aeration status, and this source was fairly constant across the sites (\u003Ca href=\"#F9\"\u003EFigure 9\u003C\u002Fa\u003E). The exception to this was at the aerators when they were ON; there was a large methane sink (negative value in \u003Ca href=\"#F9\"\u003EFigure 9\u003C\u002Fa\u003E). However, since the atmospheric flux during the ON status is mostly likely underestimated, this exception is most likely a methane source too.\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 9\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g009.jpg\" name=\"Figure9\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g009.gif\" id=\"F9\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 9\u003C\u002Fstrong\u003E. Water column methane inventory anomaly assuming the sediments are the only source of methane to the water and the air-water interface is the only sink. A positive value means that there must be a source of methane to balance the source and sink, whereas the negative value means there must be a sink consuming methane.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Ch3 class=\"pt0\"\u003E3.6 Time-Series Water Column Methane Measurements\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EUsing OsmoSamplers, two separate records of dissolved methane concentrations from bottom water were obtained. The first was from the OsmoSampler deployment in 2018 that spanned 1-week using fast pumps with ∼1 day resolution (\u003Ca href=\"#F10\"\u003EFigure 10\u003C\u002Fa\u003E). The time stamps assigned were verified by comparing salinity (as calculated from chloride concentrations) in the OsmoSampler coils and the salinity from the sensor packages (\u003Ca href=\"#SM1\"\u003ESupplementary Figure S6\u003C\u002Fa\u003E). The salinity comparison shows relatively good agreement, especially at station RC7. The overall trend is similar between the sensor and OsmoSamplers at RC1 and RC2, but as we have observed in previous studies, the absolute salinity values did not match well at these stations (\u003Ca href=\"#B24\"\u003EGelesh et al., 2016\u003C\u002Fa\u003E). The highest methane concentrations came 3 days after the aerators were turned OFF at RC2 and reached almost 3,000 nM (\u003Ca href=\"#F10\"\u003EFigure 10\u003C\u002Fa\u003E). Concentrations were also high at RC1 during this time. The timing of this methane peak came right after an event where dissolved oxygen (measured between stations RC2, and RC7) increased rapidly to ∼8 mg L\u003Csup\u003E−1\u003C\u002Fsup\u003E oxygen (\u003Ca href=\"#F10\"\u003EFigure 10\u003C\u002Fa\u003E). After 14 July 2018, methane concentrations decreased to less than 1,000 nM and were similar at all stations. Methane concentrations from OsmoSamplers were cross-checked with our discrete samples and we see no statistical difference between the two (\u003Cem\u003Ep\u003C\u002Fem\u003E = 0.72); which is the first time this has been verified in field tests. Methane concentrations at RC7 remained lower than the other stations.\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 10\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g010.jpg\" name=\"Figure10\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g010.gif\" id=\"F10\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 10\u003C\u002Fstrong\u003E. Methane concentrations in high temporal resolution in bottom water from fast OsmoSamplers (black and white symbols) and discrete water samples (red symbols) in 2018 from stations RC1 (filled stars), RC2 (open stars), and RC7 (filled circles). Discrete samples overlap with OsmoSampler concentrations. Thin black line shows a dissolved oxygen record from sensors deployed in the bottom water off a nearby dock.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cp class=\"mb0\"\u003EThe second time-series record of methane concentration came from OsmoSamplers deployed through the summer and into the fall of 2018 and 2019, and contained slow pumps that give ∼ weekly resolution (\u003Ca href=\"#F11\"\u003EFigure 11\u003C\u002Fa\u003E). The temporal pattern was not the same each year. In 2018, at RC1 and RC2, the initial concentrations before aeration were lower (∼400 nM), and then almost doubled when aerators were turned OFF (\u003Ca href=\"#F11\"\u003EFigure 11A\u003C\u002Fa\u003E). Once they were turned back ON after our experiment, concentrations at RC1 continued to decrease at a rate of ∼13 nM day\u003Csup\u003E−1\u003C\u002Fsup\u003E (linear fit with \u003Cem\u003ER\u003C\u002Fem\u003E\u003Csup\u003E2\u003C\u002Fsup\u003E = 0.8); whereas at RC2, concentrations continued to increase through July and finally peak in August at 17,000 nM. Concentrations at RC2 then decreased and reached ∼1,000 nM for the remainder of the timeseries. At RC7, concentrations didn’t show much change with time and averaged 842 ± 265 nM. In 2019, we captured much higher temporal resolution with the samplers which started about 2 weeks before our experiment began (\u003Ca href=\"#F11\"\u003EFigure 11B\u003C\u002Fa\u003E). Overall, concentrations were lower than in 2018 and ranged between 110 and 1,667 nM. There were concentration differences across sites, where methane concentrations at RC1 averaged 368 ± 100 nM; RC2 averaged 558 ± 136 nM; and RC7 averaged 400 ± 270 nM (\u003Ca href=\"#F11\"\u003EFigure 11B\u003C\u002Fa\u003E). The bottom water temperature varied between 23–28°C in 2018 with some variability (\u003Ca href=\"#SM1\"\u003ESupplmentary Figure S7\u003C\u002Fa\u003E), whereas in 2019, the temperature gradually increased from ∼23°C to a high of ∼30°C in August and then decreased into the fall where a sudden decreased to less than 20°C occurred when the aerators were turned off (\u003Ca href=\"#SM1\"\u003ESupplementary Figure S7\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003EFIGURE 11\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-g011.jpg\" name=\"Figure11\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-g011.gif\" id=\"F11\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003EFIGURE 11\u003C\u002Fstrong\u003E. Methane concentrations in bottom water in \u003Cstrong\u003E(A)\u003C\u002Fstrong\u003E 2018, and \u003Cstrong\u003E(B)\u003C\u002Fstrong\u003E 2019. Shaded region indicates when the aerators were turned off, otherwise, they were on.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Ca id=\"h5\" name=\"h5\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003E4 Discussion\u003C\u002Fh2\u003E\u003Cp class=\"mb15\"\u003EOne solution to estuarine eutrophication is to artificially aerate the waters with bubble systems. This solution has benefits for reintroducing oxygen back into the water, but it could also have consequences for methane cycling. Previous studies have documented that the hypoxic or anoxic conditions in bottom waters that result from eutrophication also lead to the build-up of dissolved methane diffusing into the bottom waters from the sediments (\u003Ca href=\"#B3\"\u003EBange et al., 2010\u003C\u002Fa\u003E; \u003Ca href=\"#B24\"\u003EGelesh et al., 2016\u003C\u002Fa\u003E) which can result in a greater atmospheric flux. Thus, we hypothesized that when aerators are placed in such a system, the physical movement of all that water, with the fact that methane has low solubility, would enhance an atmospheric methane flux. To our knowledge, there is only one other study in a temperate lake that has studied oxygen effects on methane dynamics, and they found a remarkable decrease in methane build-up with engineered aeration (\u003Ca href=\"#B32\"\u003EHounshell et al., 2021\u003C\u002Fa\u003E), yet they did not quantify air-water flux. With our dataset, we were able to directly calculate this flux when the aerators were ON versus OFF to determine the impact of aeration in terms of methane dynamics. In addition to this, we also considered that the addition of oxygen to the water column might stimulate microbial aerobic methane oxidation which would somewhat control the release of methane at the air-water interface. Through our whole ecosystem manipulation experiment, we were able to address the following questions: 1) what is the effect of aeration on the atmospheric methane flux, 2) is Rock Creek an atmospheric methane source, 3) what is the source of water column methane in the Rock Creek, and 4) is aerobic methane oxidation enhanced in the water column? We also gained insights into complex biogeochemical processes and potential feedbacks occurring in this sub-estuary during and after aeration that sharpens our focus for future studies to further elucidate critical mechanisms related to dissolved oxygen dynamics and associated biogeochemical effects.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E4.1 Aeration Enhanced Atmospheric Methane Flux\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EWe hypothesized that the air-water methane flux would be higher during aeration then when the aerators were OFF. When we simply apply the stagnant boundary layer model to calculate the fluxes, we see that sites within the aeration zone (RC1 and RC2) had higher methane fluxes than downstream, regardless of aeration status (\u003Ca href=\"#F8\"\u003EFigure 8\u003C\u002Fa\u003E). Yet these fluxes are most likely underestimates when the aerators are ON, as we see with the modified flux calculation (\u003Ca href=\"#F8\"\u003EFigure 8\u003C\u002Fa\u003E, extended arrows to blue dots). We further hypothesized that under small bubble aeration, the flux would be lower than under large bubble aeration (\u003Ca href=\"#F1\"\u003EFigure 1\u003C\u002Fa\u003E); which is what the fluxes showed in 2018 (large bubble aeration) versus 2019 (fine bubble aeration, \u003Ca href=\"#F8\"\u003EFigure 8\u003C\u002Fa\u003E), supporting this hypothesis. The other observation was that in 2019, when the fine bubble aeration was installed, the air-water methane flux was lower during aeration then when the aerators were turned OFF. We have already stated that this flux is most likely underestimated. Future work would benefit from directly measuring this flux with floating chambers to more precisely quantify this flux.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E4.2 Is Rock Creek an Atmospheric Methane Source?\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003ERock Creek is a source of atmospheric methane, regardless of aeration status, or site. The air-water methane flux from Rock Creek varied between 0.2 and 1.5 mmol m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E (note change in units to compare to literature values), which was similar to fluxes measured from several estuaries (\u003Ca href=\"#T3\"\u003ETable 3\u003C\u002Fa\u003E), and higher than those from oceanic environments, which vary between 0.0001 and 0.1 mmol m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E (\u003Ca href=\"#B4\"\u003EBižić et al., 2020\u003C\u002Fa\u003E). Rock Creek methane fluxes are on par with a shallow subarctic lake which reached almost 0.4 mmol m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E (\u003Ca href=\"#B34\"\u003EJansen et al., 2020\u003C\u002Fa\u003E), even though at times of ice-out, these lakes can release as much as 75 mmol m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E (\u003Ca href=\"#B57\"\u003EMcIntosh Marcek et al., 2021\u003C\u002Fa\u003E). Surface water concentrations were also similar to those measured from an aerated eutrophic lake (\u003Ca href=\"#B51\"\u003EMartinez and Anderson, 2013\u003C\u002Fa\u003E) suggesting methane is emitted in these aerated waters. A unique aspect to the work presented here is the high-frequency sampling over the warm season in Rock Creek which measured consistently high concentrations (400–1,000 nM) in the bottom water (\u003Ca href=\"#F11\"\u003EFigure 11\u003C\u002Fa\u003E) that rival what has been measured in the anoxic bottom waters of the mainstem Chesapeake Bay in mid-summer (\u003Ca href=\"#B24\"\u003EGelesh et al., 2016\u003C\u002Fa\u003E) and further supports a sustained methane flux to the atmosphere. Thus this relatively shallow (∼3 m) eutrophic estuary, may contribute more methane than previously thought, as was the case for streams and rivers (\u003Ca href=\"#B76\"\u003EStanley et al., 2016\u003C\u002Fa\u003E), and conforms to our understanding of coastal ecosystems as having an outsized influence on methane fluxes in a global context.\u003C\u002Fp\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"Imageheaders\"\u003ETABLE 3\u003C\u002Fdiv\u003E\u003Cdiv class=\"FigureDesc\"\u003E\u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_m\u002Ffenvs-10-866152-t003.jpg\" name=\"Table3\" target=\"_blank\"\u003E\u003Cimg src=\"https:\u002F\u002Fwww.frontiersin.org\u002Ffiles\u002FArticles\u002F866152\u002Ffenvs-10-866152-HTML-r1\u002Fimage_t\u002Ffenvs-10-866152-t003.gif\" id=\"T3\" alt=\"www.frontiersin.org\"\u002F\u003E\u003C\u002Fa\u003E\u003Cp\u003E\u003Cstrong\u003ETABLE 3\u003C\u002Fstrong\u003E. Examples of estuarine flux of methane to the atmosphere.\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"DottedLine\"\u003E\u003C\u002Fdiv\u003E\u003Ch3 class=\"pt0\"\u003E4.3 The Source of Methane: All From Sediments?\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003ESedimentary methanogenesis is likely the main source of methane to the water column of Rock Creek because the highest dissolved methane concentrations were measured in the sediments, and methane concentrations in the bottom water were typically higher than the surface water. Biogenic methane is also supported with the sedimentary porewater methane δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values in the deep sediments being < −70‰ (\u003Ca href=\"#B81\"\u003EWhiticar, 1999\u003C\u002Fa\u003E). Yet, there was also evidence that methane produced in the deep sediments went through some degree of microbial oxidation before reaching the overlying water. The sediment porewater methane profiles showed classic concave-up shapes which are indicative of the anaerobic oxidation of methane (AOM) working in concert with sulfate reduction, as expressed here with the sulfate methane transition (SMT) depths (\u003Ca href=\"#B35\"\u003EJørgensen et al., 2019\u003C\u002Fa\u003E). AOM is also supported with the porewater methane isotopic composition data. In 2019, the porewater δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values also increased up the core through the SMT depth. This pattern indicates AOM; as the methane diffuses along the concentration gradient, microbial communities preferentially utilize \u003Csup\u003E12\u003C\u002Fsup\u003EC and leave the \u003Csup\u003E13\u003C\u002Fsup\u003EC behind (thereby values increase) as methane is oxidized (\u003Ca href=\"#B81\"\u003EWhiticar, 1999\u003C\u002Fa\u003E). In 2019, this pattern is clear regardless of aeration status but there is a shift to more \u003Csup\u003E13\u003C\u002Fsup\u003EC depleted values when aerators were OFF. This shift could represent enhanced microbial methane production when aerators were OFF but would need to be validated with other information such as diagenetic modeling (e.g., \u003Ca href=\"#B50\"\u003EMartens et al., 1998\u003C\u002Fa\u003E). The 2019 sedimentary profiles, measured under the low-turbulence diffuse system, support our classic understanding of biogeochemical zonation (\u003Ca href=\"#B23\"\u003EFroelich et al., 1979\u003C\u002Fa\u003E) and suggests that the sediments are diffusion dominated.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EThe pattern in 2018 was not as clear, possibly due to sediment disturbance with vigorous aeration. The destratification system employed during 2018 involved a high-volume through flow of air that leads to substantial physical disturbance of the water-column and sediments. Such disturbance at the aerators translated into variable porewater methane concentration and the isotope patterns, compared to outside the aeration zone (\u003Ca href=\"#F5\"\u003EFigure 5\u003C\u002Fa\u003E). This implies that a simple, steady-state 1D, diffusion dominated interpretation of these profiles cannot be applied here because the destratification system could have driven substantial advective exchange between sediments and the water-column. It is interesting to note that at the aerators (RC1 and RC2), there was a depletion of \u003Csup\u003E13\u003C\u002Fsup\u003EC up the core to the sediment-water interface. One way to inject such a depleted signature is by methanogenesis which could happen in this agitated system by bubbling out deep methane during aeration or methanogenesis in the surface layers using non-competitive substrates (\u003Ca href=\"#B2\"\u003EAlperin et al., 1988\u003C\u002Fa\u003E). More work would be needed to support or refute these possibilities.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EMethane from the sediments can either diffuse into the overlying water or bubble out \u003Cem\u003Evia\u003C\u002Fem\u003E ebullition (methane oversaturated porewaters forming bubbles). The methane released from both of those processes would have different isotopic signatures. For example, for bubbles to form in the sediments and efflux, the methane concentration must be above saturation which occurs in the sediments deeper than ∼10 cmbsf, depending on the site and year (see \u003Ca href=\"#F5\"\u003EFigures 5\u003C\u002Fa\u003E, \u003Ca href=\"#F6\"\u003E6\u003C\u002Fa\u003E). At these depths, the δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values were between −80 and −85‰, so we would expect to see these values in the overlying water column (assuming a small amount of methane from the bubble equilibrates with the water). Bubbles captured in a shallow water column off North Carolina showed no isotopic fractionation when released to the water (\u003Ca href=\"#B13\"\u003EChanton and Martens, 1988\u003C\u002Fa\u003E). However, if the methane is diffusing into the water from the sediment surface, that methane would carry a δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E value similar to the surface sediment methane signature. For 2018, the δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values of water column methane were always higher than the surface sediment, regardless of site or aeration status. This suggests that there could be some methane oxidation at the very sediment surface that we are missing in our sediment measurements. This was the same situation when the aerators were OFF in 2019. Perhaps future work could focus on the sediment-water interface as an area of intense methane oxidation, be it either aerobic or anaerobic.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EThe situation is a little different when the aerators were ON in 2019; the water column methane always had a lower δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E value than the surficial sediments. This suggests that the methane was not simply diffusing in from the sediments, and that there must be another source of methane injecting isotopically depleted carbon; such as, sediment ebullition, methane being advected out of the sediments into the water column from the aerators, or microbial methane production in the water column. We don’t have data to support or refute the advective release of methane out of the sediments other than to say that in 2019, the aeration was much finer and thus most likely less advective than in 2018 so it seems unlikely. Plus, since we were not focused on determining the ebullitive flux, its hard to evaluate. However, in 2018, we had evidence of an ebullitive flux of methane from the sediments. First, the methane in the air above Rock Creek waters had an average δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E value of −50‰ which is slightly depleted in \u003Csup\u003E13\u003C\u002Fsup\u003EC from the global, well-mixed value from the northern hemisphere of −47.4‰ (\u003Ca href=\"#B40\"\u003ELan et al., 2021\u003C\u002Fa\u003E). When this air δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E value was compared to what was measured in the deep sediments, < −70‰, or in the water column, < −60‰, the \u003Csup\u003E13\u003C\u002Fsup\u003EC depleted methane in air over Rock Creek waters can be explained with a small amount of biogenic methane from the sediments. Secondly, we measured a direct pulse of biogenic (δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E = −80‰) methane in the air above RC7 at dawn when the aerators were OFF (\u003Ca href=\"#F7\"\u003EFigure 7\u003C\u002Fa\u003E). This was most likely due to ebullition from the sediments directly reaching the atmosphere which was trapped in the air above the water due to the air inversion that occurs at dawn (\u003Ca href=\"#B16\"\u003ECrill et al., 1988\u003C\u002Fa\u003E; \u003Ca href=\"#B60\"\u003EMukhophadhya et al., 2001\u003C\u002Fa\u003E). While this observation of high concentrations of methane in the morning air is not unexpected, it clearly documents how aqueous environments contribute to the atmospheric methane isotopic signal which has recently been shown to be enhanced in biogenic methane sources (\u003Ca href=\"#B73\"\u003ESchaefer et al., 2016\u003C\u002Fa\u003E; \u003Ca href=\"#B40\"\u003ELan et al., 2021\u003C\u002Fa\u003E). It also clearly shows that in order to fully capture methane dynamics in this system, future work should quantify the bubble flux from the tributary, much like was done in the temperature lake in California (\u003Ca href=\"#B51\"\u003EMartinez and Anderson, 2013\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E4.4 Evidence for Methane Production in the Water Column\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EWhile we might have captured methane bubbling out of Rock Creek in 2018, that process is very heterogeneous and not continuous, and may not fully explain the water column δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values from 2019 ON status. We thus explore the possibility of water column methane production under aeration. There is growing evidence for an oxic production pathway in surface waters (see review in \u003Ca href=\"#B4\"\u003EBižić et al., 2020\u003C\u002Fa\u003E) that might be important. Most notably, it was concluded that 90% of methane in surface waters of a temperate lake was formed in the oxic surface waters, and not the sediments (\u003Ca href=\"#B21\"\u003EDonis et al., 2017\u003C\u002Fa\u003E). However, a recent reevaluation of this work concluded that a sedimentary methane flux from the sediments flanking the lake in the shallow waters could explain the oversaturated surface waters (\u003Ca href=\"#B64\"\u003EPeeters et al., 2019\u003C\u002Fa\u003E). In well stratified lakes, methane production in surface waters is shown to scale with sediment area and mixed layer volume (\u003Ca href=\"#B26\"\u003EGünthel et al., 2019\u003C\u002Fa\u003E). Lakes are hydrodynamically very different from estuaries, making study comparisons to Rock Creek tenuous. Furthermore, the hydrodynamic conditions in destratification systems, such as in 2018, can create large cells of overturning water that may impact a much larger area of the tidal system (\u003Ca href=\"#B25\"\u003EGibbs and Howard-Williams, 2018\u003C\u002Fa\u003E). The lack of studies of sub-estuaries alone highlights the need to quantify the sources of methane from systems such as Rock Creek.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EThe box model, which balances sedimentary and air-water methane fluxes, shows that for most of the time, regardless of if aerators were ON or OFF, there is an additional methane source in the water column in 2019 (\u003Ca href=\"#F11\"\u003EFigure 11\u003C\u002Fa\u003E). The observation that outside the aeration zone, RC7 and RC9, there is also an additional methane source could indicate the transport of methane from the Patapsco River, lateral inputs of water from across the creek as destratification cells pull in sources from a cross-section of the creek or simply upstream methane sources. Previous work in rivers also documented higher methane flux in upstream surface waters (\u003Ca href=\"#B18\"\u003Ede Angelis and Scranton, 1993\u003C\u002Fa\u003E; \u003Ca href=\"#B1\"\u003EAbril and Iversen, 2002\u003C\u002Fa\u003E; \u003Ca href=\"#B58\"\u003EMiddelburg et al., 2002\u003C\u002Fa\u003E); as well as higher methane in smaller width creeks (\u003Ca href=\"#B5\"\u003EBorges and Abril, 2011\u003C\u002Fa\u003E).\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EThe further observation that there is a methane source regardless of the status of the aerators (either ON or OFF) suggests there is a bigger ecosystem response than just the influx of oxygen from the aerators. One possibility could be the presence of algal blooms which produce methane as a byproduct (\u003Ca href=\"#B4\"\u003EBižić et al., 2020\u003C\u002Fa\u003E). Such blooms have been shown to occur before when aerators are turned OFF (\u003Ca href=\"#B28\"\u003EHarris et al., 2015\u003C\u002Fa\u003E), and we also documented one right after the aerators were turned OFF in 2018 (\u003Ca href=\"#F10\"\u003EFigure 10\u003C\u002Fa\u003E). Increased surface water oxygen levels were observed with an increase in particulate organic nitrogen and carbon, and iron (data not shown). The OsmoSamplers also captured a pulse of methane right after this oxygen pulse. If these blooms are a typical phenomenon right after turning the aerators OFF, it could explain why the methane fluxes were higher 3 days after turning aerators OFF in 2021 (\u003Ca href=\"#F4\"\u003EFigure 4D\u003C\u002Fa\u003E) than when the aerators were ON. Future work could also directly quantify rates of methane production in aerobic waters.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E4.5 Evidence for Aerobic Methane Oxidation in the Water Column?\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EWhile it seems clear that the aerators are enhancing a methane flux to the air above the creek, that the sediments are the main source of methane to the waters and there possibly is methane production in these aerated waters, we also considered if there is any microbially mediated oxidation in the surface waters that might offset this atmospheric flux. The idea here is that the waters are being oxygenated due to the aeration, as the oxygen data suggested happened during the experiment (\u003Ca href=\"#F3\"\u003EFigure 3\u003C\u002Fa\u003E), and this oxygen is allowing for aerobic methane oxidation to happen. We have already speculated about a small potential oxidation process altering the δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E from our surficial sedimentary methane values to the bottom water. Here, we focus solely on the water column. Since we didn’t directly measure microbial rates, we need to rely on geochemical data. We did this in four ways. First, we interrogated the stable carbon isotopes of methane measured in the water column. The δ\u003Csup\u003E13\u003C\u002Fsup\u003EC-CH\u003Csub\u003E4\u003C\u002Fsub\u003E values give a bulk measure of what has happened to that methane since it was formed; the bulk methane signature would be more enriched in \u003Csup\u003E13\u003C\u002Fsup\u003EC if it had been oxidized. In order to do this, we considered that the methane in the bottom water is the source for the methane in the surface water. \u003Ca href=\"#F5\"\u003EFigures 5\u003C\u002Fa\u003E, \u003Ca href=\"#F6\"\u003E6\u003C\u002Fa\u003E water column isotope profiles clearly show that in all cases, the surface water is always slightly \u003Csup\u003E13\u003C\u002Fsup\u003EC depleted, if at all different, contrary to what would be found if aerobic methane oxidation was occurring. Secondly, we looked for a correlation between oxygen concentrations and methane concentrations. For this correlation, we surmised that if oxygen was controlling methane levels, we would see a linear relationship between the two. There was no correlation (\u003Ca href=\"#SM1\"\u003ESupplementary Figure S8A\u003C\u002Fa\u003E), which suggests oxygen is not the limiting factor for aerobic methane oxidation. Thirdly, we considered that maybe aerobic methane oxidation was limited by dissolved inorganic nitrogen (DIN), as has been proposed in the literature (\u003Ca href=\"#B71\"\u003ESansone and Martens, 1978\u003C\u002Fa\u003E). We surmise that if there is a negative relationship between DIN and water column methane concentrations, there could be evidence for aerobic methane oxidation. In other words, if methanotrophs were stimulated by DIN in an otherwise oxic water column, we would find low methane concentrations. This pattern did not emerge (\u003Ca href=\"#SM1\"\u003ESupplementary Figure S8B\u003C\u002Fa\u003E). And finally, in the box model approach described in Section 4.3, for most of the time in 2018 and 2019, there was a source of methane and not a sink. The exception was in 2019 when the aerators were ON. Taking all of this geochemical evidence into consideration, we conclude that we found little to no geochemical evidence for water column aerobic methane oxidation. Future work could focus on looking for microbial signature of oxidation in the water column.\u003C\u002Fp\u003E\u003Ch3 class=\"pt0\"\u003E4.6 Complicating Factors With Experimental Design\u003C\u002Fh3\u003E\u003Cp class=\"mb0\"\u003EWe have discussed the differences in atmospheric methane flux between the 2018 and 2019 aeration experiments to most likely be a factor of the bubble size during aeration. The large bubble aerators resulted in a larger flux of methane to the atmosphere than the small bubble aerators. However there were also differences between the 2 years in how long the aerators were turned OFF. Despite variability in the concentrations, we found almost no effect of aeration when the aerators were turned OFF for a few days (2016, 2021), but there was modest methane accumulation as oxygen was depleted for 3–7 days in 2018 and significant methane increases after 13 days in 2019. Thus, it appears that methane accumulation requires more than a day to emerge after deoxygenation, despite prior studies in Arctic lake systems (\u003Ca href=\"#B57\"\u003EMcIntosh Marcek et al., 2021\u003C\u002Fa\u003E) and in the Chesapeake Bay mainstem (\u003Ca href=\"#B24\"\u003EGelesh et al., 2016\u003C\u002Fa\u003E) that appear to indicate that methane is immediately released from sediments following deoxygenation. However, the temporal resolution of prior measures likely cannot capture day to day dynamics, and there are few, if any real-time measures of methane and oxygen in coastal systems to confirm the time-scale of methane responses to deoxygenation.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EA second source of complication in our interpretations is that the experimental study is embedded in a system with background environmental variability. Thus, our “ON-OFF” study design using whole ecosystem manipulations does not offer a true controlled experimental system. For example, unlike previous work in Rock Creek (\u003Ca href=\"#B28\"\u003EHarris et al., 2015\u003C\u002Fa\u003E), the waters did not go anoxic within 1 day in 2018 and 2019 and only at RC2 in 2018 did anoxia ever emerge. In 2018, there was also an influx of oxygen-rich water that appeared in the aeration zone within days after the aerators were turned OFF and when oxygen should have been depleted (seen in time series data on \u003Ca href=\"#F10\"\u003EFigure 10\u003C\u002Fa\u003E). Although we did not measure methane during this event, the rapid increase in methane concentrations immediately after this pulse of tidally driven, oxygen-rich water could be the result of a rapid response to deoxygenation (that was not interrupted by the event) or an influx of methane rich water from upstream after the event. Our measurements in 2021 did indicate higher methane concentrations upstream of the aeration zone (∼1,000 nM; \u003Ca href=\"#F4\"\u003EFigure 4D\u003C\u002Fa\u003E), but these are substantially smaller than the concentrations measured after the event in 2018 (1,500 to 2,500 nM; \u003Ca href=\"#F10\"\u003EFigure 10\u003C\u002Fa\u003E), suggesting that an upstream source is unlikely. Other potential inputs of methane are currently unknown, such as groundwater, and there are no substantial tidal wetlands that could serve as a methane source. Although groundwater remains unlikely since chloride concentrations slightly increased with depth in sediment porewaters, our understanding of the drivers for the patterns remains elusive and long term observatories are needed.\u003C\u002Fp\u003E\u003Cp class=\"mb0\"\u003EEcosystem responses to aeration beyond deoxygenation also likely feedback to influence methane dynamics in the estuary. \u003Ca href=\"#B28\"\u003EHarris et al. (2015)\u003C\u002Fa\u003E reported a substantial algal bloom in Rock Creek within a week of the aerators being turned off, generating high surface water oxygen and organic matter concentrations. Given that Rock Creek has substantial light attenuation (Secchi Depth < 0.5), the vertical mixing induced by aerators likely keeps phytoplankton mixed below the photic layer, and when aerators are turned OFF, water-column stabilization allows for phytoplankton blooms. This is consistent with recent work in Chesapeake Bay that suggest that high chlorophyll-a packaging under low-light conditions combined with high nutrient concentrations allows for rapid algal growth when light becomes available (\u003Ca href=\"#B8\"\u003EBuchanan, 2020\u003C\u002Fa\u003E). Methane concentrations reached the highest levels we measured at RC2 in 2018 (∼2,700 nM), 3-days after an increase in particulate organic carbon of 700 μM after the aerators were turned OFF (data not shown). The consumption of this organic matter that likely followed may have generated substantial new methane, especially considering that oxygen concentrations were consistently below 32 μM (1 mg L\u003Csup\u003E−1\u003C\u002Fsup\u003E) for much of the following week (\u003Ca href=\"#F10\"\u003EFigure 10\u003C\u002Fa\u003E). Given that the “destratification” approach used in 2018 was designed to physically mix and overturn the waters, such a phytoplankton response is likely. The “diffuse” approach used in 2019 involved much less physical disruption of the water-column (water-column salinity profiles were similar during ON and OFF), which may have prevented a phytoplankton response and also allowed for stable conditions that allowed the microbial communities to organize along expected redox conditions. This finding also points to an important impact of aerator design on our results, with differences in impact dependent on the mechanism of aeration and questions remain regarding how best to quantify mixing, oxygenation, and even water fluxes dependent on whether engineering impacted circulation (destratification design through large bubbles) versus diffusion of oxygen (small bubble aeration design).\u003C\u002Fp\u003E\u003Ca id=\"h6\" name=\"h6\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003E5 Conclusion\u003C\u002Fh2\u003E\u003Cp class=\"mb15\"\u003EThere is evidence that the sediments are the main source of methane in Rock Creek as conceptualized in \u003Ca href=\"#F1\"\u003EFigure 1\u003C\u002Fa\u003E, although we cannot rule out upstream creek waters or \u003Cem\u003Ein situ\u003C\u002Fem\u003E production in the water column as additional sources. Our measurements suggest that aeration decreases the time frame available for aerobic methane oxidation in the water-column, thus connecting the sediment to the atmosphere more directly. This study did not allow for an in-depth examination of seasonal variations in methane sources and sinks, nor did it measure the impact of other complex factors on methane, such as shifting nutrient status and related microbial responses. Our experiments were also not intended to test aeration system design and operation (bubble size and density) in relation to methane flux, instead we used models to evaluate these impacts. Future work should 1) couple methane concentration and isotope data along with microbial rate measurements under various nutrient and oxygen conditions upstream and downstream of the aerators; 2) investigate seasonal changes to understand the complex factors controlling methane flux from this eutrophic estuary; and 3) characterize aerator design impacts on local hydrodynamics to better characterize physical effects on sediment transport and sediment-water exchange.\u003C\u002Fp\u003E\u003Cp class=\"mb15\"\u003EKeeping these recommendations in mind, the strength of our study can be distilled into two central findings we emphasize here: 1) The shallow, Rock Creek sub-estuary is a source of methane to the atmosphere, regardless of engineering intervention, and methane production and flux is likely enhanced as a consequence of eutrophication and 2) the strength of the methane flux is impacted by aeration bubble size design. Our conceptual model lays out the processes and impacts that are connected to these findings (\u003Ca href=\"#F1\"\u003EFigure 1\u003C\u002Fa\u003E). The smallest fluxes occurred with the small bubble system (\u003Ca href=\"#F9\"\u003EFigure 9\u003C\u002Fa\u003E), and we predict that continued implementation of this system combined with potential future oligotrophication could reduce methane fluxes. The largest fluxes occurred with the older, large bubble system that was intended to encourage destratification. The data support the hypothesis that aeration can lead to higher atmospheric methane fluxes and that aerator design is crucial to mitigating methane transfer. A key motivation for this study was to investigate the potential for unintended consequences of this engineering intervention in relation to greenhouse gas emissions. The dependency of the measured methane atmospheric fluxes on bubble size suggests that there is a path forward towards optimizing aerator design to reduce this consequence in eutrophic tidal waters (e.g., implement small bubble aeration). This study also adds to a growing body of literature quantifying methane fluxes in coastal waters. Regardless of aerator status or design, the current condition of Rock Creek as a eutrophic ecosystem characterized by high primary production of organic matter impacts its overall role as a source of methane to the atmosphere. As coastal water quality policies are implemented and managers seek both solutions and greater understanding of the complex biogeochemistry that impacts restoration trajectories in eutrophic systems, work on both engineering solutions and interpretation of restoration monitoring data would benefit from including methane dynamics and greenhouse gas impacts into holistic management frameworks.\u003C\u002Fp\u003E\u003Ca id=\"h7\" name=\"h7\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003EData Availability Statement\u003C\u002Fh2\u003E\u003Cp class=\"mb15\"\u003EThe datasets presented in this study can be found in online repositories. The names of the repository\u002Frepositories and accession number(s) can be found below: \u003Ca href=\"https:\u002F\u002Fwww.ncei.noaa.gov\u002Farchive\u002Faccession\u002F0244510\"\u003Ehttps:\u002F\u002Fwww.ncei.noaa.gov\u002Farchive\u002Faccession\u002F0244510\u003C\u002Fa\u003E.\u003C\u002Fp\u003E\u003Ca id=\"h8\" name=\"h8\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003EAuthor Contributions\u003C\u002Fh2\u003E\u003Cp class=\"mb0\"\u003ELL, AH, JT, and LH conceived the idea; LL, CH, MF, CS, and EH conducted the experiment and analyzed samples; LL wrote the initial manuscript, and all authors contributed significant edits to the manuscript. We thank Mark Mobley for help with the bubble model, and the reviewers and editors who greatly improved this manuscript.\u003C\u002Fp\u003E\u003Ca id=\"h9\" name=\"h9\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003EFunding\u003C\u002Fh2\u003E\u003Cp class=\"mb0\"\u003EFinancial support for this project was through U.S. National Science Foundation CBET-1706416 (LH, AH, JT, and LL), Maryland Sea Grant (NA14OAR4170090, SA75281450-O; LH and JT), and Maryland SeaGrant REU program (NSF grant OCE-1262374).\u003C\u002Fp\u003E\u003Ca id=\"h10\" name=\"h10\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003EConflict of Interest\u003C\u002Fh2\u003E\u003Cp class=\"mb0\"\u003EThe authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.\u003C\u002Fp\u003E\u003Ca id=\"h11\" name=\"h11\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003EPublisher’s Note\u003C\u002Fh2\u003E\u003Cp class=\"mb15\"\u003EAll claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. Any product that may be evaluated in this article, or claim that may be made by its manufacturer, is not guaranteed or endorsed by the publisher.\u003C\u002Fp\u003E\u003Ca id=\"h12\" name=\"h12\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003EAcknowledgments\u003C\u002Fh2\u003E\u003Cp class=\"mb03\"\u003EWe thank Anne Arundel county for access to the aerators, the residents of Rock Creek for allowing us to conduct the experiments in their backyard, and Janis Markusic for her support. The work could not have been done without the field help of Zachary Gotthardt, Lillian Henderson, and Brittany Clark and laboratory help of Maureen Strauss and Hadley McIntosh Marcek. We thank Dave Oliff at the Florida State University machine shop for building the fast OsmoSamplers. Work on this paper by Lapham was possible with the expertise and care of healthcare workers at the Johns Hopkins Breast Cancer Center and support from Marcia Lapham and Dr. Cédric Magen. Co-author Harris was supported by the Rock Creek team as a nursing mother in the field with patience, privacy, and good humor during boat work with co-authors Szewczyk and Testa. Lapham, Harris, and Testa depended on childcare provided by their communities and especially their spouses during the COVID-19 pandemic and are thankful to have had the time for this research. This is UMCES contribution #6188.\u003C\u002Fp\u003E\u003Ca id=\"h13\" name=\"h13\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003ESupplementary Material\u003C\u002Fh2\u003E\u003Cp class=\"mb15\" id=\"SM1\"\u003EThe Supplementary Material for this article can be found online at: \u003Ca href=\"https:\u002F\u002Fwww.frontiersin.org\u002Farticles\u002F10.3389\u002Ffenvs.2022.866152\u002Ffull#supplementary-material\"\u003Ehttps:\u002F\u002Fwww.frontiersin.org\u002Farticles\u002F10.3389\u002Ffenvs.2022.866152\u002Ffull#supplementary-material\u003C\u002Fa\u003E\u003C\u002Fp\u003E\u003Ca id=\"h14\" name=\"h14\"\u003E\u003C\u002Fa\u003E\u003Ch2\u003EReferences\u003C\u002Fh2\u003E\u003Cdiv class=\"References\"\u003E\u003Cp class=\"ReferencesCopy1\"\u003E\u003Ca name=\"B1\" id=\"B1\"\u003E\u003C\u002Fa\u003EAbril, G., and Iversen, N. (2002). Methane Dynamics in a Shallow Non-tidal Estuary (Randers Fjord, Denmark). \u003Cem\u003EMar. Ecol. Prog. 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Chem.\u003C\u002Fem\u003E 172, 12–22. doi:10.1016\u002Fj.marchem.2015.03.004\u003C\u002Fp\u003E\u003Cp class=\"ReferencesCopy2\"\u003E\u003Ca href=\"https:\u002F\u002Fdoi.org\u002F10.1016\u002Fj.marchem.2015.03.004\"\u003ECrossRef Full Text\u003C\u002Fa\u003E | \u003Ca href=\"https:\u002F\u002Fscholar.google.com\u002Fscholar?hl=en&as_sdt=0%2C5&q=Observing+Methane+Hydrate+Dissolution+Rates+Under+Sediment+Cover&btnG=\"\u003EGoogle Scholar\u003C\u002Fa\u003E\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"References\"\u003E\u003Cp class=\"ReferencesCopy1\"\u003E\u003Ca name=\"B84\" id=\"B84\"\u003E\u003C\u002Fa\u003EYamamoto, S., Alcauskas, J. B., and Crozier, T. E. (1976). Solubility of Methane in Distilled Water and Seawater. \u003Cem\u003EJ. Chem. Eng. Data\u003C\u002Fem\u003E 21 (1), 78–80. doi:10.1021\u002Fje60068a029\u003C\u002Fp\u003E\u003Cp class=\"ReferencesCopy2\"\u003E\u003Ca href=\"https:\u002F\u002Fdoi.org\u002F10.1021\u002Fje60068a029\"\u003ECrossRef Full Text\u003C\u002Fa\u003E | \u003Ca href=\"https:\u002F\u002Fscholar.google.com\u002Fscholar?hl=en&as_sdt=0%2C5&q=Solubility+of+Methane+in+Distilled+Water+and+Seawater&btnG=\"\u003EGoogle Scholar\u003C\u002Fa\u003E\u003C\u002Fp\u003E\u003C\u002Fdiv\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"thinLineM20\"\u003E\u003C\u002Fdiv\u003E\u003Cdiv class=\"AbstractSummary\"\u003E\u003Cp\u003E\u003Cspan\u003EKeywords:\u003C\u002Fspan\u003E methane, aeration, eutrophication, estuary, OsmoSampler\u003C\u002Fp\u003E\u003Cp\u003E\u003Cspan\u003ECitation:\u003C\u002Fspan\u003E Lapham LL, Hobbs EA, Testa JM, Heyes A, Forsyth MK, Hodgkins C, Szewczyk C and Harris LA (2022) The Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary. \u003Cem\u003EFront. Environ. Sci.\u003C\u002Fem\u003E 10:866152. doi: 10.3389\u002Ffenvs.2022.866152\u003C\u002Fp\u003E\u003Cp id=\"timestamps\"\u003E\u003Cspan\u003EReceived:\u003C\u002Fspan\u003E 30 January 2022; \u003Cspan\u003EAccepted:\u003C\u002Fspan\u003E 13 June 2022;\u003Cbr\u002F\u003E\u003Cspan\u003EPublished:\u003C\u002Fspan\u003E 11 August 2022.\u003C\u002Fp\u003E\u003Cdiv\u003E\u003Cp\u003EEdited by:\u003C\u002Fp\u003E \u003Ca href=\"https:\u002F\u002Floop.frontiersin.org\u002Fpeople\u002F606160\u002Foverview\"\u003EDaniel F. McGinnis\u003C\u002Fa\u003E, Université de Genève, Switzerland\u003C\u002Fdiv\u003E\u003Cdiv\u003E\u003Cp\u003EReviewed by:\u003C\u002Fp\u003E \u003Ca href=\"https:\u002F\u002Floop.frontiersin.org\u002Fpeople\u002F1668703\u002Foverview\"\u003EPeter Tango\u003C\u002Fa\u003E, United States Geological Survey (USGS), United States\u003Cbr\u002F\u003E\u003Ca href=\"https:\u002F\u002Floop.frontiersin.org\u002Fpeople\u002F218446\u002Foverview\"\u003EWei-Dong Zhai\u003C\u002Fa\u003E, Shandong University, China\u003C\u002Fdiv\u003E\u003Cp\u003E\u003Cspan\u003ECopyright\u003C\u002Fspan\u003E © 2022 Lapham, Hobbs, Testa, Heyes, Forsyth, Hodgkins, Szewczyk and Harris. This is an open-access article distributed under the terms of the \u003Ca rel=\"license\" href=\"http:\u002F\u002Fcreativecommons.org\u002Flicenses\u002Fby\u002F4.0\u002F\" target=\"_blank\"\u003ECreative Commons Attribution License (CC BY).\u003C\u002Fa\u003E The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.\u003C\u002Fp\u003E\u003Cp\u003E\u003Cspan\u003E*Correspondence:\u003C\u002Fspan\u003E Laura L. Lapham, \u003Ca href=\"mailto:lapham@umces.edu\"\u003Elapham@umces.edu\u003C\u002Fa\u003E\u003C\u002Fp\u003E\u003Cdiv class=\"clear\"\u003E\u003C\u002Fdiv\u003E\u003C\u002Fdiv\u003E",menuHtml:"\u003Cul class=\"flyoutJournal\"\u003E\u003Cli\u003E\u003Ca href=\"#h1\"\u003EAbstract\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h2\"\u003E1 Introduction\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h3\"\u003E2 Materials and Methods\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h4\"\u003E3 Results\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h5\"\u003E4 Discussion\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h6\"\u003E5 Conclusion\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h7\"\u003EData Availability Statement\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h8\"\u003EAuthor Contributions\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h9\"\u003EFunding\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h10\"\u003EConflict of Interest\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h11\"\u003EPublisher’s Note\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h12\"\u003EAcknowledgments\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h13\"\u003ESupplementary Material\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003Cli\u003E\u003Ca href=\"#h14\"\u003EReferences\u003C\u002Fa\u003E\u003C\u002Fli\u003E\u003C\u002Ful\u003E"},files:[{name:"EPUB.epub",fileServerPackageEntryId:f,type:{code:ax,name:ax}},{name:ay,fileServerPackageEntryId:"fenvs-10-866152-r1\u002Ffenvs-10-866152.pdf",type:{code:t,name:t}},{name:ay,fileServerPackageEntryId:f,type:{code:t,name:t}},{name:"fenvs-10-866152.xml",fileServerPackageEntryId:"fenvs-10-866152-r1\u002Ffenvs-10-866152.xml",type:{code:"NLM_XML",name:"XML"}},{name:"Provisional 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Laboratory, United States","Frontiers in Environmental Science","PDF",5,627,"environmental-science",4,2,"description","Frontiers","Help center","Link","Grey","Medium","ssph-journal.org","fship","Front. Environ. Sci.","2296-665X",void 0,18,1920,"por-journal.com",7,"escubed.org",1918,"fipp","https:\u002F\u002Fd2csxpduxe849s.cloudfront.net\u002Fmedia\u002FE32629C6-9347-4F84-81FEAEF7BFA342B3\u002FC5D37CA5-E038-46DD-8CD87C9172B5B36F\u002Fwebimage-E986D64D-768A-4D21-970E7D73635C3DA2.png","image","2022-06-27T09:59:36Z","fenvs",25,"journal_field","10.3389\u002Ffenvs.2022.866152","The Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary","\u003Cp\u003EEngineered aeration is one solution for increasing oxygen concentrations in highly eutrophic estuaries that undergo seasonal hypoxia. Although there are various designs for engineered aeration, all approaches involve either destratification of the water column or direct injection of oxygen or air through fine bubble diffusion. To date, the effect of either approach on estuarine methane dynamics remains unknown. Here we tested the hypotheses that 1) bubble aeration will strip the water of methane and enhance the air-water methane flux to the atmosphere and 2) the addition of oxygen to the water column will enhance aerobic methane oxidation in the water column and potentially offset the air-water methane flux. These hypotheses were tested in Rock Creek, Maryland, a shallow-water sub-estuary to the Chesapeake Bay, using controlled, ecosystem-scale deoxygenation experiments where the water column and sediments were sampled in mid-summer, when aerators were ON, and then 1, 3, 7, and 13 days after the aerators were turned OFF. Experiments were performed under two system designs, large bubble and fine bubble approaches, using the same observational approach that combined discrete water sampling, long term water samplers (OsmoSamplers) and sediment porewater profiles. Regardless of aeration status, methane concentrations reached as high as 1,500 nmol L\u003Csup\u003E−1\u003C\u002Fsup\u003E in the water column during the experiments and remained near 1,000 nmol L\u003Csup\u003E−1\u003C\u002Fsup\u003E through the summer and into the fall. Since these concentrations are above atmospheric equilibrium of 3 nmol L\u003Csup\u003E−1\u003C\u002Fsup\u003E, these data establish the sub-estuary as a source of methane to the atmosphere, with a maximum atmospheric flux as high as 1,500 µmol m\u003Csup\u003E−2\u003C\u002Fsup\u003E d\u003Csup\u003E−1\u003C\u002Fsup\u003E, which is comparable to fluxes estimated for other estuaries. Air-water methane fluxes were higher when the aerators were ON, over short time frames, supporting the hypothesis that aeration enhanced the atmospheric methane flux. The fine-bubble approach showed lower air-water methane fluxes compared to the larger bubble, destratification system. We found that the primary source of the methane was the sediments, however, \u003Citalic\u003Ein situ\u003C\u002Fitalic\u003E methane production or an upstream methane source could not be ruled out. Overall, our measurements of methane concentrations were consistently high in all times and locations, supporting consistent methane flux to the atmosphere.\u003C\u002Fp\u003E",24,23188,"Laura L.",1413210,"Edward A.",435564,"Jeremy M.",1675968,"Andrew","Melinda K.","Casey",1710495,"Curtis",1264726,"Lora A.",606160,218446,"Wei-dong",1668703,"Peter",{},1436,"Biogeochemical Dynamics","biogeochemical-dynamics",10,"EPUB","fenvs-10-866152.pdf","Frontiers | The Effects of Engineered Aeration on Atmospheric Methane Flux From a Chesapeake Bay Tidal Tributary","https:\u002F\u002Fwww.frontiersin.org\u002Fjournals\u002Fenvironmental-science\u002Farticles\u002F10.3389\u002Ffenvs.2022.866152\u002Ffull","Engineered aeration is one solution for increasing oxygen concentrations in highly eutrophic estuaries that undergo seasonal hypoxia. 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