205
15 Biogeochemical Indicators of Nutrient Enrichments in Wetlands …
and associated nutrient cycling. The final section is devoted
to the evaluation of the model of microbial allocation of
resources among community indicator enzymes (MARCIE
model) as a tool for wetland eutrophication assessment
using microbial indicators.
15.2 Summary of the Main Biogeochemical
Changes Associated with Nutrient
Enrichment in Wetlands
The first step of nutrient enrichment in wetlands is an increasing productivity of wetland plants. High primary productivity associated with nutrient-rich areas has been observed in numerous wetlands (Davis 1991; Childers et al.
2003; Álvarez-Cobelas et al. 2010). It promotes increases
of autochthonous organic matter accumulation in wetland
soils (Sánchez-Carrillo and Álvarez-Cobelas 2001). Usually,
an increase of nutrient loading is linked to hydrological alterations and shifts in vegetation patterns can be expected
(i.e., plant species replacement; Green and Galatowitsch
2002; Childers et al. 2003; Álvarez-Cobelas et al. 2008).
Therefore, changes in organic matter supply and litter quality can be expected in high nutrient scenarios (Engelhardt
and Ritchie 2002). Because microbial activity responds to
nutrient loading and hydrology (DeBusk and Reddy 1998;
Wright and Reddy 2001), also strong shifts in decomposition rates and nutrient cycling can be expected as wetland
become eutrophic.
After prolonged exposure to high nutrient loading an increase of the wetland internal loading can be expected and
hence a reduction in nutrient retention occurs. Because sediment suspension into the water column tend to increase the
nutrient soluble forms during eutrophication (Reddy et al.
1998), more available nutrients promote more productivity,
resulting in a positive feedback process which can last for
years. Thus, nutrient loading into these wetlands enhances
organic matter decomposition and microbial activity which
increases nutrient concentrations in water (Wright et al.
2008). Although decomposition in these wetlands appear to
be dominated by anaerobic pathways (denitrification, sulfate
reduction, and methanogenesis; Wright and Reddy 2001),
aerobic heterotrophic microbial activity is often enhanced by
nutrient loading too (Aerts and Toet 1997; Qualls and Richardson 2000). In the Everglades it has been observed that
heterotrophic microorganisms are limited by the high C:P
ratios of organic matter but external P loading remove this
limitation and induce changes in microbial activity (DeBusk
and Reddy 1998). However, aerobic CO 2 production could
not be related to the content of dissolved nutrients in soils
(Wright et al. 2009) but with microbial biomass, which appeared positively related to nutrient loading (D’Angelo and
Reddy 1999). While heterotrophic microbial activity is immediately enhanced after the addition of oxygen in nutrientrich wetland experiments, net soil organic carbon mineralization has not been shown significant differences when other
electron acceptors as NO 3
− , SO 4
2− , or CO 2 under anaerobic
conditions has been used (D’Angelo and Reddy 1999).
Microbial biomass phosphorus (hereafter MBP) in wetland soils responds positively to phosphorus enrichment in
wetlands (Qualls and Richarson 2000; Newman et al. 2003).
Labile phosphorus (the most available P fraction to vegetation and microbial communities; Ivanoff et al. 1998) has
been cited as the most sensitive P fraction to eutrophication
in wetland soils and seston (consisted of algae, periphyton
and particulate organic matter; White et al. 2006). Microbial
biomass C and N (hereafter MBC and MBN) did not consistently respond to nutrient loading (Wright et al. 2008, 2009).
MBC and MBN are linked to C and N cycles but cannot be
considered sensitive to nutrient loading. Although eutrophication promotes mineralization rates, phosphorous mineralization has been cited as the most responsive microbial indicator to nutrient enrichment in wetlands (Wright et al. 2009).
Increased availability of sulfate in wetlands is known to
cause serious eutrophication problems, as S
2−
produced by
SO 4
2− reduction interacts with Fe-PO 4
3− complexes in the sediment to produce FeS 2 and FeS, resulting in mobilization of
phosphate (Smolders and Roelofs 1993; Lamers et al. 1998).
High nitrate concentrations in ground waters have been cited
to inhibit eutrophication of sulfate-rich freshwater wetlands
(Lucassen et al. 2004). Since NO 3
− is an energetically more
favorable electron acceptor in anaerobic wetland soils than
Fe
3+
and SO 4
2− , high NO 3
− loads function as a redox buffer,
preventing the reduction of Fe
3+
and SO 4
2− . Therefore, limited SO 4
2− reduction prevents S
2−
-mediated mobilization of
PO 4
3− from Fe-PO 4
3− complexes. At higher redox potential,
reduced Fe
2+
is oxidized, increasing the content of Fe
3+
capable to binding PO 4
3− (Reddy and DeLaune 2008).
While the typical loading rates of nitrogen and phosphorus in natural and constructed freshwater wetlands exceed
proposed critical loads to prevent eutrophication (total phosphorus: 10 kg P ha
−1
y
−1
, total nitrogen: 25 kg N ha
−1
y
−1
),
wetlands continue to have serious problems preserving its
structure (species composition) and functioning (nutrient
cycling and retention; Verhoeven et al. 2006). The potential
impact of climate change on wetlands will promote ecosystem changes because temperature and precipitation are
strong determinants of freshwater wetland ecosystem structure and function (Mitsch and Gosselink 2000). Lower water
availability will reduce the flushing rate of wetlands as water
renewal times increases. It will probably contribute to increase the high nutrient state of wetlands as biogeochemical
processes will be promoted. In the same way, because wetlands are a major natural source of greenhouse gases to the
atmosphere we may expect changes in the emissions of N 2 O
and N 2 as a consequence of eutrophication. Nitrous oxide
15 Biogeochemical Indicators of Nutrient Enrichments in Wetlands …
and associated nutrient cycling. The final section is devoted
to the evaluation of the model of microbial allocation of
resources among community indicator enzymes (MARCIE
model) as a tool for wetland eutrophication assessment
using microbial indicators.
15.2 Summary of the Main Biogeochemical
Changes Associated with Nutrient
Enrichment in Wetlands
The first step of nutrient enrichment in wetlands is an increasing productivity of wetland plants. High primary productivity associated with nutrient-rich areas has been observed in numerous wetlands (Davis 1991; Childers et al.
2003; Álvarez-Cobelas et al. 2010). It promotes increases
of autochthonous organic matter accumulation in wetland
soils (Sánchez-Carrillo and Álvarez-Cobelas 2001). Usually,
an increase of nutrient loading is linked to hydrological alterations and shifts in vegetation patterns can be expected
(i.e., plant species replacement; Green and Galatowitsch
2002; Childers et al. 2003; Álvarez-Cobelas et al. 2008).
Therefore, changes in organic matter supply and litter quality can be expected in high nutrient scenarios (Engelhardt
and Ritchie 2002). Because microbial activity responds to
nutrient loading and hydrology (DeBusk and Reddy 1998;
Wright and Reddy 2001), also strong shifts in decomposition rates and nutrient cycling can be expected as wetland
become eutrophic.
After prolonged exposure to high nutrient loading an increase of the wetland internal loading can be expected and
hence a reduction in nutrient retention occurs. Because sediment suspension into the water column tend to increase the
nutrient soluble forms during eutrophication (Reddy et al.
1998), more available nutrients promote more productivity,
resulting in a positive feedback process which can last for
years. Thus, nutrient loading into these wetlands enhances
organic matter decomposition and microbial activity which
increases nutrient concentrations in water (Wright et al.
2008). Although decomposition in these wetlands appear to
be dominated by anaerobic pathways (denitrification, sulfate
reduction, and methanogenesis; Wright and Reddy 2001),
aerobic heterotrophic microbial activity is often enhanced by
nutrient loading too (Aerts and Toet 1997; Qualls and Richardson 2000). In the Everglades it has been observed that
heterotrophic microorganisms are limited by the high C:P
ratios of organic matter but external P loading remove this
limitation and induce changes in microbial activity (DeBusk
and Reddy 1998). However, aerobic CO 2 production could
not be related to the content of dissolved nutrients in soils
(Wright et al. 2009) but with microbial biomass, which appeared positively related to nutrient loading (D’Angelo and
Reddy 1999). While heterotrophic microbial activity is immediately enhanced after the addition of oxygen in nutrientrich wetland experiments, net soil organic carbon mineralization has not been shown significant differences when other
electron acceptors as NO 3
− , SO 4
2− , or CO 2 under anaerobic
conditions has been used (D’Angelo and Reddy 1999).
Microbial biomass phosphorus (hereafter MBP) in wetland soils responds positively to phosphorus enrichment in
wetlands (Qualls and Richarson 2000; Newman et al. 2003).
Labile phosphorus (the most available P fraction to vegetation and microbial communities; Ivanoff et al. 1998) has
been cited as the most sensitive P fraction to eutrophication
in wetland soils and seston (consisted of algae, periphyton
and particulate organic matter; White et al. 2006). Microbial
biomass C and N (hereafter MBC and MBN) did not consistently respond to nutrient loading (Wright et al. 2008, 2009).
MBC and MBN are linked to C and N cycles but cannot be
considered sensitive to nutrient loading. Although eutrophication promotes mineralization rates, phosphorous mineralization has been cited as the most responsive microbial indicator to nutrient enrichment in wetlands (Wright et al. 2009).
Increased availability of sulfate in wetlands is known to
cause serious eutrophication problems, as S
2−
produced by
SO 4
2− reduction interacts with Fe-PO 4
3− complexes in the sediment to produce FeS 2 and FeS, resulting in mobilization of
phosphate (Smolders and Roelofs 1993; Lamers et al. 1998).
High nitrate concentrations in ground waters have been cited
to inhibit eutrophication of sulfate-rich freshwater wetlands
(Lucassen et al. 2004). Since NO 3
− is an energetically more
favorable electron acceptor in anaerobic wetland soils than
Fe
3+
and SO 4
2− , high NO 3
− loads function as a redox buffer,
preventing the reduction of Fe
3+
and SO 4
2− . Therefore, limited SO 4
2− reduction prevents S
2−
-mediated mobilization of
PO 4
3− from Fe-PO 4
3− complexes. At higher redox potential,
reduced Fe
2+
is oxidized, increasing the content of Fe
3+
capable to binding PO 4
3− (Reddy and DeLaune 2008).
While the typical loading rates of nitrogen and phosphorus in natural and constructed freshwater wetlands exceed
proposed critical loads to prevent eutrophication (total phosphorus: 10 kg P ha
−1
y
−1
, total nitrogen: 25 kg N ha
−1
y
−1
),
wetlands continue to have serious problems preserving its
structure (species composition) and functioning (nutrient
cycling and retention; Verhoeven et al. 2006). The potential
impact of climate change on wetlands will promote ecosystem changes because temperature and precipitation are
strong determinants of freshwater wetland ecosystem structure and function (Mitsch and Gosselink 2000). Lower water
availability will reduce the flushing rate of wetlands as water
renewal times increases. It will probably contribute to increase the high nutrient state of wetlands as biogeochemical
processes will be promoted. In the same way, because wetlands are a major natural source of greenhouse gases to the
atmosphere we may expect changes in the emissions of N 2 O
and N 2 as a consequence of eutrophication. Nitrous oxide